Food Safety Issues for Mineral and Organic Fertilizers

C H A P T E R T W O

Food Safety Issues for Mineral and Organic Fertilizers Rufus L. Chaney Contents 1. Introduction 2. Concentrations of Trace Elements in Fertilizers and Soil Amendments 3. Risk Assessment Pathways 4. Natural Controls on Trace Element Transfer from Soil to Plants 5. Element Risks 5.1. Selenium 5.2. Molybdenum 5.3. Cobalt 5.4. Cadmium 5.5. Lead 6. Trace Elements in Other Common Soil Amendments 6.1. Manure 6.2. Limestone 6.3. Steel production fume waste 6.4. Gypsum 7. Long-Term Reactions of Trace Elements in Soils 8. Other Elements in Fertilizers and Soil Amendments of Possible Concern 8.1. Fluoride 8.2. Radionuclides 8.3. Arsenic 9. Monitoring and Control of Trace Elements in Mineral Fertilizers 10. Need for Regulatory Enforcement on Composition of Soil Amendments References

52 53 57 58 59 59 60 61 62 83 84 84 85 86 86 87 89 89 92 94 97 98 99

Senior research Agronomist, USDA-Agricultural Research Service, Environmental Management and Byproducts Utilization Lab, BARC-West, Beltsville, MD, USA Advances in Agronomy, Volume 117 ISSN 0065-2113, DOI: http://dx.doi.org/10.1016/B978-0-12-394278-4.00002-7

2012 Published by Elsevier Inc. All rights reserved.

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Rufus L. Chaney

Abstract Fertilizers and other soil amendments are required to maintain soil fertility, but some may be naturally rich in trace elements, or contaminated. As part of the overall consideration of using fertilizers and soil amendments, one should consider the levels of trace elements present in relation to soil, plant, and food-chain processes (precipitation, adsorption, chelation) which promote or alleviate trace element risks. These natural processes limit plant accumulation of nearly all elements to levels which would not cause harm to humans, livestock, wildlife, or soil organisms. Soils geologically rich or contaminated with Mo can harm ruminants, while those rich in Se may harm all plant consumers; Mo or Se should be applied only when needed. Manures from swine and poultry may be rich in Cu, Zn and/or As from feed additives. Crops except rice accumulate little As from soils, so soil ingestion is the basis for soil As risk except for rice. Pb risk is also through soil ingestion rather than plant uptake. Cd is accumulated by rice to levels which caused human disease (renal tubular dysfunction) where rice soils were contaminated by industrial discharges. Risk from Cd in rice is strongly affected by the high bioavailability of rice Cd. Consumption of similar amounts of Cd have not caused harm from other foods. Because phosphate fertilizers may contain high levels of Cd, and use of high Cd superphosphate in Australia caused significant increase in wheat and potato Cd levels, risk from long-term accumulation of phosphate fertilizer Cd (and other sources) must be controlled. Different control schemes are discussed.

1. Introduction Cases of excessive food-chain transfer of soil or applied trace elements indicate that all fertilizers and soil amendments require evaluation. Some products may be contaminated even when nearly all similar products are quite low in trace elements. Thus, control systems are required to assure compliance with quality standards to protect purchasers and food safety. On the other hand, soil and crop properties have prevented transfer of excessive trace elements in nearly all situations, so details are needed to understand why some elements and products have caused adverse effects in some cases while most do not. Historically, veterinary toxicology has shown that certain elements in soil or amendments can poison livestock if the element is applied in excess, or occurs naturally at high levels in the soil. In particular, Se and Mo have been known to harm livestock grazing mineralized alkaline soils for longer than 100 years (McDowell, 2003; Underwood, 1977). More recently, CdeZn mine waste contamination of rice soils in Asia caused excessive accumulation of Cd but not Zn in rice grain, such that subsistence farm families suffered Cd health effects (Kobayashi, 1978; Nogawa et al., 2004). Because Cd was long known to occur in phosphate fertilizers

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(e.g., Williams and David, 1973), these findings in Asia caused concern about food-chain transfer of fertilizer-applied Cd. But research has helped explain why significant Cd disease has only been found in subsistence rice consumers with industrially-contaminated paddy soils (Chaney, 2010). Great concern has been expressed about some fertilizer products which contain high levels of Cd, Pb, and As, although nearly all As and Pb concern is about garden fertilizer products which may be ingested by children rather than garden crop uptake of As and Pb. More recently, excessive ingestion of sulfate from well water, soil amendments, or feed additives has caused a sulfur toxicity disease in ruminants (Gould, 1998). These examples of identified problems from excessive food-chain transfer of trace elements in agriculture raised many questions about how the diseases occurred, and how better management of fertilizers and soil amendments could prevent future problems. Much research has been conducted on these issues for decades, giving us a better understanding of the soileplanteanimal relationships of trace elements. In addition, improved science supported risk assessment by government agencies that sorted out which elements and products need controls, and which should be handled by management practices. On the other hand, excessive concern based on poor science could lead to unnecessarily restrictive limits on trace elements in fertilizers, and raise the costs of these products and of foods to consumers. This paper does not consider biosolids use in agriculture. I have reviewed the topic of biosolids and biosolids composts in several recent publications (Chaney and Ryan, 1993; Chaney and Ryan, 1994; Chaney et al., 2001; Basta et al., 2005). It should be recognized that intense research on biosolids has supported the development of understanding of science about soil and amendment trace element risk assessment (US-EPA, 1995).

2. Concentrations of Trace Elements in Fertilizers and Soil Amendments Specific fertilizer and soil amendment products (phosphates; byproduct Zn fertilizers) normally contain significant levels of elements of potential concern (Cd, As, Pb, F, etc.), while others (N and K fertilizers) normally contain only very low levels of these elements (Table 1). Table 2 shows that phosphates from different sources contain a wide range of Cd concentrations depending on the rock phosphate source and processing technology (McLaughlin et al., 1996; Chen et al., 2007; Grant et al., 1999). Among soil amendments other than pesticides, arsenic is

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Table 1

Arsenic and cadmium concentrations in commercial fertilizers in California As Mean

Cd Median

P Fertilizers Rock Phosphates Multi-nutrient Fertilizers Biosolids Fertilizers Zn Micronutrient Sulfur Micronutrient Iron Micronutrient ZneFeeMn Blends

Mean

Median

89 33 37 2.8 149 0.43 71 248

132 0.5 19 1.0 23 0 21 e

Range

11 7.4 13 2.4 30 5.8 1544 48

13 7.5 8.0 1.7 0.5 1.0 118 e

0e21 1.4e13 0.15e155 0e10 0e280 0.1e19 0.3e4950 24e71

0e163 0e130 0e200 0e15 0e495 0e3.0 0e334 95e400 Rufus L. Chaney

(CDFA, 1998).

Range mg kgL1

Fertilizer Type

Composition and limits for Cd in P fertilizers in several countries

Cd:P Country

Limits for Fertilizer-Cd Washington USA-Oregon USA-California AAPFCO Australia Canada Japan Austria Belgium Denmark Netherlands Finland Sweden EU Proposal (2001)

Limit

0.0889 kg Cd ha
1 yr
1 7.5 mg Cd (%P2O5)
1 4 mg Cd >(%P2O5)
1 10 mg Cd (%P2O5)
1 300 mg Cd (kg P)
1 0.0889 kg Cd ha
1 yr
1 75 mg Cd (kg P2O5)
1 90 mg Cd (kg P2O5)
1 21.5 mg Cd (kg
1 P2O5) 43 mg Cd (kg
1 P2O5) 20 mg Cd kg
1 P2O5 40 mg Cd kg
1 P2O5 60 mg Cd kg
1 P2O5

mg Cd kg

L1

2040 774 412 1030 300 2040 340 275 206 110 40 49 100 45.8 91.6 137.0

P

Cd:P2O5

Cd in 45% P2O5-Product

mg Cd kgL1 P2O5

mg Cd kgL1

889 338 180 450 131 889 148 120 90 48.0 17.5 21.5 43.7

400 152 81 202 59 400 67 54 40.5 21.6 7.9 9.7 19.7

20 40 60

9.0 18 27

Food Safety Issues for Mineral and Organic Fertilizers

Table 2

(continued ) 55

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Table 2 (continued ) Cd:P L1

Limit

mg Cd kg

Levels present in example products Kola USA-FL Sweden USA-NC Nauru High-Western-US

1.2 46 118 336 641 1170

Country

P

Cd:P2O5

Cd in 45% P2O5-Product

mg Cd kgL1 P2O5

mg Cd kgL1

0.523 20 51 147 280 511

0.24 9 23 (TSP) 66 126 230

(Grant et al., 1999; data for Swedish traditional phosphates from Andersson and Hahlin, 1981).

Rufus L. Chaney

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only above background soil concentrations in a few phosphate products and some livestock manures. But phosphate products produced from most ores contain Cd above the levels in soils except for electric furnace phosphate products. The concentrations are expressed in different ways because different groups conducting risk assessment focus on Cd:P ratio, Cd:P2O5 ratio, or concentration of Cd in the actual product. There is general acceptance that limits for Cd or other elements in P-fertilizers should be expressed per unit P or P2O5. Chien and Menon (1994) compared rock phosphate, acidulated rock phosphate, and normal phosphate fertilizers and stressed that one should also consider Cd on the basis of Cd per unit phytoavailable P; P phytoavailability varies widely among rock phosphates as both P and Cd have varied solubility, but the Cd in rock phosphate eventually becomes labile according to several studies (Kuo et al., 2007).

3. Risk Assessment Pathways In order to evaluate whether trace elements in specific fertilizers or soil amendments might comprise risk, one must evaluate the potential for the element to cause adverse effects to organisms be they humans, livestock, wildlife, or soil microbes. Over the last few decades, pathways for assessment of potential risks have become established during the risk assessments for land application of numerous materials. The pathways to humans were always clear: the general agricultural pathway focused on grains; the garden foods pathway focused on fruits and vegetables grown in home gardens; the livestock pathways to humans; water, air, and soilderived dust. Livestock and wildlife health are assessed both for plants grown on amended soils, and for ingestion of the amended soil or the amendment itself adhering on plant leaves. Wildlife have different pathways than humans and livestock, and bioaccumulation may play a greater role in wildlife food-chains than for humans; because earthworms accumulate Cd within their bodies, earthworm consuming wildlife become a highly-exposed species for soil Cd enrichment. The pathways have been well described and used in risk assessments in the US and the EU (US-EPA, 1995). Careful evaluation of all the pathways for trace element transfers and risks allows appropriate assessment of which levels of contaminants are required to protect food safety and the environment. But appropriate assessment requires evaluation of interactions which limit element transfer and risk, not just the worst possible cases which are often stressed by those seeking more strict controls.

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4. Natural Controls on Trace Element Transfer from Soil to Plants Natural properties of elements, soils, plants, and animals greatly limit the potential for risk from nearly all elements in a process I have called the “Soil-Plant Barrier” (Chaney, 1983; Langmuir et al., 2005; McLaughlin et al., 1999). Most elements are so insoluble in soils or in roots that they do not reach the edible shoots of plants at levels which could cause even chronic toxicity to plants or consumers (e.g., Cr3þ, Hg, F, Pb, Ag, Ba, Zr, V, rare earths, etc.). Other elements may be taken up and readily translocated to edible plant tissues, but phytotoxicity limits the maximum concentration in edible plant tissues with visible injury to levels which are chronically tolerated by livestock, wildlife, and humans (Zn, Cu, Ni, Mn, F, As, B). Reduced yield or death of the plant limits exposure of consumers to these elements in soil. Livestock are usually the most exposed and susceptible plant consumers because they may consume 100% of their diet as plants which are experiencing significant yield reduction due to phytotoxicity. And ruminants are sensitive to some elements (Mo, Se, S) for which non-ruminants are much more tolerant. A few elements are readily absorbed and translocated to plant shoots without causing phytotoxicity. These are the elements which are known to potentially cause food-chain toxicity (Mo, Se, Cd, and possibly Co). These elements are considered in some detail herein. Another process allows exposure of animals to elements in soils, the direct ingestion of soil or soil amendment. Grazing livestock are known to consume appreciable amounts of soil either as splash on leaves, or over-grazed pastures, or as amendments adhering to plant leaves (e.g., Chaney and Lloyd, 1979). And young children ingest housedust (partly soil which has been carried into homes by parents, siblings or pets) by hand-to-mouth play. Risk assessment research has found that young children ingest an average of 24  4 (SD) mg soil d
1 with the 95th percentile ingestion rate of 91  16.6 mg d
1 (Stanek et al., 2001). Even if elements such as Pb, As, or F are not accumulated by plants to dangerous levels because of the chemistry of soils and plants, the soil element may be sufficiently bioavailable in ingested soil to cause adverse effects to livestock, wildlife, or even humans. Elements for which soil or soil amendment ingestion is the principal source of potential risk to humans or livestock include: Pb, As, F, Fe2þ, Cr6þ and Hg. Many elements occur in fertilizers and soil amendments at measurable levels, but in most products the elements are at low concentrations and cannot cause food-chain contamination issues even after centuries of use. But elements in some materials require controls to prevent the identified

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potential problems. All products must be protected from deliberate contamination. Understanding why and how some elements need to be controlled in fertilizers and soil amendments helps with understanding how humans, livestock and wildlife are protected against this potential risk by control systems or natural processes.

5. Element Risks 5.1. Selenium The Mo and Se poisoning of livestock are complicated, but illustrate soilplant-animal processes which allow risk. First, Se is well known as a natural toxin in geologically mineralized soils in several US states. On the one hand, most US soils are deficient in Se for human and livestock health (Kubota et al., 1987). Increasing interest in improving Se levels in human diets has resulted in tests of Se fertilization (e.g., Bañuelos, 2006; Broadley et al., 2010). Because selenate ðSeO2
4 Þ is an anion, and it is strongly adsorbed on soil hydrous Fe oxide at low soil pH, plant uptake of Se is very limited in acidic soils, but can reach high levels in alkaline soils if Se is present. The other half of Se risk through food-chain transfer is that plants are very tolerant of Se, such that plants tolerate higher levels than do the animals that consume the crops. A typical maximum chronic dietary level for livestock is 2 mg kg
1 dry weight (NRC, 1980), while forage plants can accumulate 100 mg kg
1 without harm to the plant. And there are even hyperaccumulators of Se which reach 10,000 mg Se kg
1 (Chaney et al., 2010). Selenate is usually absorbed by the sulfate transporter of roots, so if sulfate is simultaneously high in soils, selenate accumulation is strongly reduced (Stroud et al., 2010), except for the hyperaccumulators (Bell et al., 1992). Concern was raised about Se in land-applied fly ash from coal-fired power plants. If Se is present in the coal, much Se enters the fly ash collected during exhaust filtration. When plants were grown on fly ash landfills, the plants contained enough Se to poison wildlife (Arthur et al., 1992). But when similar fly ash was applied at fertilizer or limestone replacement rates, the Se only raised plant Se to levels sufficient for livestock for a few years (Gutenmann et al., 1979). Se is more dangerous in aquatic than terrestrial ecosystems because the Se can be bioaccumulated in algae and biomagnified in lower animals because it can be built into proteins as selenomethionine and selenocysteine. These problems were discovered at the Kesterson Reservoir in the Central Valley of California where Se-enriched drainage waters in evaporation ponds caused severe adverse Se effects on wildlife. The Se was leached from irrigated Se-mineralized soils which did not cause

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accumulation of excessive Se levels in forages. But when evaporation raised concentrations, and aquatic organisms accumulated and then biomagnified Se, poisoning of wildlife followed (Ohlendorf et al., 1986). These observations caused wide concern about excessive Se applications, while large numbers of humans and livestock remained with inadequate Se for best health (Kubota and Allaway, 1972; Kubota et al., 1987; Broadley et al., 2010). Thus Se fertilizer applications are limited to those required to satisfy food quality goals, and to limit potential runoff, leaching, or excessive accumulation in plants or aquatic ecosystems. Livestock diets are routinely supplemented with Se, and most of that Se ends up in manure applied on cropland. But no problems have been reported from Se in manure of normally Se-supplemented livestock manures. However, in an unusual case in China, ash from burning of Semineralized coal was applied on cropland; maize (Zea mays L.) accumulation of Se poisoned both livestock and humans (Yang et al., 1983). Observing the livestock health problems should have allowed diagnosis of the Se problem, but the location was remote and human Se toxicity disease occurred.

5.2. Molybdenum Risks from Mo have traditionally occurred where poorly-drained Momineralized soils were alkaline because this element is present as the molybdate anion ðMoO2
4 Þ (Kubota et al., 1987; O'Connor et al., 2001). Molybdate is bound strongly by Fe oxides in acidic soils, but is more soluble and readily accumulated by many plants in alkaline soils to concentrations which can poison ruminant livestock but have not been found to comprise risk to non-ruminants. In the rumen, sulfate is reduced to sulfide and reacts with Mo to form thiomolybdates. Thiomolybdates form non-bioavailable complexes with Cu in the rumen (or in the body) and induce Cu deficiency when plant tissues contain over about 10 mg Mo kg
1 dry weight. Mo toxicity is complex because the concentration and form, and hence bioavailability of Cu in plants varies, and other soil factors can inhibit Cu absorption by ruminants (O’Connor et al., 2001). Extremely high sulfate in ruminant diets can also induce Cu deficiency and other adverse effects (see below). So risk assessment for Mo is interrelated with the management of Cu and S levels of diet and water, and levels of soil ingestion which can also inhibit Cu absorption (Suttle et al., 1984). Locations with high geogenic Mo and alkaline soils often cannot be used for ruminant livestock production because of the potential for forages or grains to accumulate excessive Mo. Cu supplementation or injection can readily counteract excessive Mo ingestion when it is diagnosed.

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One example of a soil-plant-food chain Mo problem involved smelter emission of Mo in PA where some nearby soils were strongly acidic. For many years the low uptake of Mo from acidic soils prevented any adverse effects. But when the soils were limed, Mo uptake was greatly increased and cattle were harmed (Hornick et al., 1977). Neighboring farms which had used limestone application for many years to improve soil fertility had no Mo toxicity because the higher pH promoted leaching of Mo from the root zone. Plants are highly tolerant of Mo. When acidic soils require Mo fertilization to support growth of plant species with higher Mo requirements (e.g., cauliflower, legumes), it may be useful to add limestone to increase the phytoavailability of Mo instead, or in addition to Mo fertilizers so that future excessive Mo for ruminants will not occur upon liming. Legumes commonly accumulate higher levels of Mo than do grasses, so the specific crops being grown on mineralized or fertilized soils must be considered in preventing problems. A comprehensive consideration of allowable Mo levels in land-applied biosolids discusses all factors known to interact with potential Mo excess in ruminants (O’Connor et al., 2001). If biosolids containing no higher than 40 mg Mo kg
1 are annually applied at N-fertilizer rates to fertilize a mixture of forage species, no adverse effects are predicted based on numerous field tests of Mo phytoavailability from biosolids. If soils are alkaline which promotes Mo uptake, the same soil pH promotes leaching from the root zone and alleviates Mo risk over time (Phillips and Meyer, 1993). One biosolids with extreme Mo contamination from an industrial source would have caused severe Mo toxicity in ruminant livestock if it had been used for forage production (Pierzynski and Jacobs, 1986).

5.3. Cobalt The case of soil Co risk to ruminants is theoretical; no adverse effects have been observed in agricultural practice. Rumen microbes require Co for normal metabolism. But when fed with diets more than 10 mg Co kg
1 DW, young cattle can be poisoned by Co (NRC, 1980; Ely et al., 1948). This concentration is lower than the concentration required to cause visual symptoms of Co phytotoxicity in most plant species (25e50 mg kg
1), so the “Soil-Plant Barrier” does not prevent Co poisoning of ruminants (Chaney, 1983). However, few soils or soil amendments contain high levels of Co, and excessive Co in forage diets has not been reported to date. Legumes also require Co for N fixation, and some soils require Co fertilization for full yield of some legumes. So this theoretical risk from excessive Co requires attention for evaluation of Co-rich soil amendments and Co fertilizer rates. Natural soils rich in Co include serpentine soils with 100e400 mg Co kg
1, but the same soils contain about 10-fold higher Ni levels which inhibit Co uptake, and can

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cause Ni phytotoxicity before plants can attain even 10 mg Co (kg DW)
1. This interaction is illustrated in Malik et al. (2000). In other cases, some trace elements (Cu, Co) enriched P-fertilizers have adhered to leaves of forages until washed off by rainfall; grazing livestock consumed excessive trace elements and were harmed (Underwood, 1977). Co and Cu fertilizer applications for grazed pastures require attention to prevent this possible direct ingestion of the fertilizer Co and Cu. These were reported only in Australia where extensive trace element deficient pasture land has to be fertilized to support livestock production, but the livestock were not removed to other pastures during the fertilizer application.

5.4. Cadmium The last of the potentially food-chain toxic trace elements is cadmium. Although Cd can cause phytotoxicity, plants can accumulate high enough Cd concentration to comprise risk to humans before phytotoxicity is evident. Because healthy plants can accumulate enough Cd to harm humans over 50 years of chronic dietary exposure, we must limit Cd in soils and soil amendments. Adverse effects of soil Cd on human health were first observed in Japan where farm families consumed home-grown rice for decades from paddies where Cd- and Zn-rich mine wastes were deposited from upstream sources (e.g., Kobayashi, 1978). Rice grain accumulated over 1 mg Cd kg
1 (e.g., Takijima and Katsumi, 1973; Fukushima et al., 1973). The disease was called Itai-itai (ouch-ouch) because of the pain from repeated bone fractures of some in the population (about 220 of more than 16,000 with adverse effects on proximal kidney tubule function at Toyoma, Japan) (Aoshima et al., 1995; Nogawa et al., 2004). Subsequently, similar Cd kidney disease has been found in many populations consuming homegrown rice from contaminated paddy soils in Japan and China, and one in Thailand (Cai et al., 1990; Simmons et al., 2003), and bone disease was also found at one location in China ( Jin et al., 2004). Based on many studies, one can expect Cd disease in populations exposed to rice grown on soils near ZnePb mines and smelters, and any upstream Cd industry, because water-borne Cd is retained effectively in paddy soils (e.g., Simmons et al., 2005, 2009). Based on the concerns raised by the observed Cd disease in Japan, many scientists studied potential soil and food Cd risks. However, food-chain risks from Cd are very complex, and many readers have not taken this complexity into account. It has become evident that Cd in different foods has quite different bioavailability. For example, Cd in shellfish has been consumed in high amounts without causing unusual Cd absorption by humans and without harm (Sharma et al., 1983; Sirot et al., 2008; Copes

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et al., 2008). Further, individuals with home gardens on soils highly contaminated by Zn mine or smelter wastes have not been found to suffer Cd disease (Strehlow and Barltrop, 1988; Ewers et al., 1985; Sarasua et al., 1995; de Burbure et al., 2003). Zn which accompanies Cd in these contaminated soil both inhibits Cd uptake and Cd bioavailability, but also can cause severe phytotoxicity at low leaf Cd compared to Cd contamination alone, limiting Cd in lettuce (Lactuca sativa L.) leaves to lower than 5 mg kg
1 DW (Chaney and Ryan, 1994). In order to better understand the risks of Cd in foods, controlled feeding tests with humans and laboratory animals have been conducted. Many toxicologists have fed very high Cd dietary concentrations so they could see adverse effects quickly, but high dietary Cd cannot mimic the potential chronic absorption of Cd in real foods. Fox et al. (1979) discuss in detail why levels of Cd found in foods grown on Cd-enriched soils are needed for valid research on food Cd risks. Much of the available feeding test literature is simply irrelevant because the diet Cd may have been higher than total diet Zn rather than the usual 1% or less of dietary Zn in nearly all foods. More recently it was found that Cd is predominantly absorbed by the ferrous transporter in the duodenum (DMT1) (Bannon et al., 2003; Kim et al., 2007; Park et al., 2002), helping to explain why even mild iron deficiency stress caused higher Cd absorption by humans (Flanagan et al., 1978). Increased dietary Zn can also inhibit Cd absorption by competition for the transporter (e.g., Jacobs et al., 1978). Zn grown into leafy vegetables significantly reduced Cd absorption from the crop as well (McKenna et al., 1992), and Swiss chard grown on control vs. biosolids amended soils with up to 5-times higher crop Cd and Zn did not cause any increase in kidney or liver Cd in Guinea pigs fed a high fraction of chard in their diet for a long period (Chaney et al., 1978). Although Cd is considerably higher in bran and whole wheat than in white flour, Cd absorption was not higher from the bran and whole wheat (Moberg et al., 1987; Wing, 1993). These studies suggested that the fiber and phytate reduced Cd absorption from bran and whole wheat, while their other studies showed that added Fe reduced Cd retention significantly from all of the bread diets (Wing et al., 1992). With this evidence, and the understanding that polished rice is deficient in Fe, Zn, and Ca for human health, Reeves and Chaney (2008) undertook experiments to clarify how the supply of Fe, Zn, and Ca in a crop food may affect absorption of Cd from that food, using levels of Cd present in the crop normally consumed by humans. Rats fed polished rice diets with marginal levels of Fe, Zn, and Ca (not low enough to reduce the rate of gain, but low enough to induce up-regulation of transport proteins in the intestine) absorbed 10-fold more Cd than rats fed adequate levels according to standard nutritional diets for lab animals (Reeves and Chaney, 2002). But because sunflower kernels provide more Zn, Fe,

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phytate, and fiber than polished rice, absorption from sunflower kernels was significantly less than from rice in diets, and the marginal diets had less effect on Cd absorption in sunflower than rice diets (Reeves and Chaney, 2001). They then did a study to follow the fate of Cd from a single rice meal over time in rats fed marginal vs. adequate ZneFeeCa diets, finding that the marginal diets caused about 10-fold higher absorption into the duodenum, and that the Cd remained in the duodenum up to about 10 days post feeding in the marginal animals which allowed the 10-fold higher Cd release to blood and liver and subsequent slow transfer to kidney (Reeves and Chaney, 2004). With this evidence of the large difference in how rats handled Cd vs. Zn from rice diets, Reeves et al. (2005) tested whether metallothionein (MT) was involved in Cd absorption. Toxicologists have long suggested that absorbed Cd was bound by MT, a low molecular weight cysteine-rich protein found in intestine, kidney, liver, etc., and which is known to bind stored Cd in the kidney. Using mice null for the MT gene, fed marginal and adequate ZneFeeCa diets, they showed that the absence of the MT gene had no effect on Cd absorption by the animals, and no interaction with the adequacy of ZneFeeCa in the diet on Cd absorption. Thus, the long claimed important role of MT in Cd metabolism was shown to be irrelevant to food Cd risks. MT is important in preventing acute toxicity of dietary or injected Cd, but unrelated to chronic dietary Cd exposure. Strong direct evidence whether the specific food or diet markedly affects whether Cd in that food actually comprises risk has been obtained from human volunteer feeding studies. It is well known that Cd intake from smoking shows up in the blood within 24 hrs (e.g., Ellis et al., 1979). Blood indicates recent absorption of Cd from diets, so sampling individuals consuming diets with low vs. high levels of Cd intakes from specific foods can test whether such foods have bioavailable Cd. This is clearly shown in the report of Sharma et al. (1983) in which smoking increased blood Cd much more than consuming large amounts of oyster Cd. In tests with young non-smoking women in Sweden, shellfish consumers ingested 3-times higher daily Cd on average than omnivores who did not consume shellfish; but they had no higher blood Cd (Vahter et al., 1996). Shellfish are excellent sources of bioavailable Fe and Zn that reduce dietary Cd bioavailability. Similarly, volunteers consuming twice as high dietary Cd with daily sunflower kernels than peanuts had no higher blood Cd during a year-long feeding test (Reeves et al., 2001). Reeves and Chaney (2008) reviewed the evidence that different foods have quite different bioavailability of Cd in the food, indicating that bioavailability should be considered in Cd risk assessment (not presently considered in JECFA or CODEX deliberations). Other aspects of dietary Cd risk assessment are confusing to many readers, but the science is clear that diet-induced Cd disease is very unusual

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except for rice subsistence farmers over 50 years old. Chaney and Ryan (1994) summarized Cd exposures from soils in different populations vs. measures of Cd effects noting that the normal distribution of urinary indicator proteins extends well beyond the levels used by European researchers as the cut-off between normal and adverse effect levels. The excretion of b2-microglobulin (b2MG) by humans normally increases with age, such that “normal” b2MG levels in urine reach over 1000 mg g
1 creatinine for 50e60 year olds, and over 1500 mg b2MG g
1 creatinine for 60e70 year olds (Kowal and Zirkes, 1983; Chaney and Ryan, 1994). But European researchers commonly define disease starting at 300 mg b2MG g
1 creatinine. And Japanese experience clearly shows that persons with actual renal tubular dysfunction due to excessive Cd ingestion often excrete urine containing 100,000 mg b2MG g
1 creatinine (Aoshima, 1987; Ikeda et al., 2003; Tohyama et al., 1981). Further, the cut-off for normal (“not yet injured by Cd”) is for the upper 95% Confidence Interval, which presumes that those above the cut-off are affected while there is no evidence of Cd effect until much higher b2-MG is attained. In Japan, China and Thailand contaminated areas, individuals with frank diet Cd-induced renal tubular dysfunction comprise over 50% of the highly-exposed subsistence rice consuming population over 50 years of age (Cai et al., 1990; Honda et al., 2010; Horiguchi et al., 2010; Nogawa et al., 2004). While in the European studies which claim to show adverse effects of Cd on the population at low Cd intakes have at most a few percent of the population (Åkesson et al., 2005; Alfven et al., 2000; Järup and Alfvén. 2004; Olsson et al., 2002; Suwazono et al., 2006), likely the fraction of the normal population above the 95% Confidence Interval. And northern Europeans have among of the lowest food Cd intake levels, population kidney Cd concentration (Olsson et al., 2002), and kidney cortex Cd levels (Friis et al., 1998) among developed nations. Claims from the European Food Safety Agency (EFSA, 2009, 2010) that the allowable daily intake of Cd should be reduced to 2.5 mg (kg body weight)
1 wk
1 were rejected by JECFA (2010) which slightly reduced their long standing recommendation of 7 mg Cd (kg BW)
1 wk
1 to 25 mg Cd (kg BW)
1 month
1 (30 days) (stressing the chronic aspect of Cd risk requiring high consumption of foods containing higher than normal levels of bioavailable Cd for decades before adverse effects could occur). The EFSA Opinion is contraindicated by human study results in Japan where middle-aged non-smoking urban women with much higher average Cd intake than their counterparts in Europe have no evidence of Cd disease even though the high end of their normal urinary Cd excretion is 3 mg g
1 creatinine (Ikeda et al., 2006). Another key misunderstanding of Cd risk is whether the higher consumption of Cd per unit body weight by children is fully accounted

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for in the JECFA evaluation. Some scientists stress concern about the risk from daily Cd intake by children compared to the JECFA Potentially Tolerable Weekly Intake (PTWI), but such concern is inappropriate for the lifetime chronic risk from dietary Cd. Children normally consume as much as double the Cd daily intake in terms of mg Cd (kg BW)
1 d
1 compared with adults consuming the same foods because they are growing and require more food intake per unit body weight than adults. But the model for 50 year chronic exposure required to reach the maximum acceptable kidney cortex Cd concentration is based on the 50 year food consumption pattern of normal foods and food Cd levels of a normal Swedish population (Friberg et al., 1985; Kjellström and Nordberg, 1978). Children's higher Cd consumption rate is fully included in present JECFA guidance based on chronic adult diet Cd intake. The evidence about actual rice Cd-induced renal tubular dysfunction, and the rare bone disease in those with severe Cd-induced renal tubular dysfunction, coupled with the evidence about low Cd bioavailability from shellfish and sunflower kernels compared with that of rice, indicate that consumption of excessive bioavailable Cd in western diets is much more difficult to achieve than in rice diets of subsistence farmers. Appropriate consideration of allowable safe chronic daily Cd consumption is required in order to select safe Cd levels in crops, safe soil pH management, and safe levels of Cd in agricultural amendments. CODEX established a limit for Cd in rice grain at 0.4 mg Cd kg
1 fresh weight (CODEX, 2006), but maintained the limit for wheat at 0.2 mg Cd kg
1. This is irrational in consideration of the large volume of data showing that rice comprises far higher potential for Cd risk than wheat. And in the case of durum wheat, much of the whole grain Cd is removed with the bran during milling to make durum flour. But the durum wheat is shipped as whole grain to prevent spoilage, so the whole grain Cd limit unnecessarily limits importation of durum wheat which will produce low Cd foods. Further, while many persons consume locally grown rice grain where the whole crop could have been grown on contaminated soils, few if any individuals consume only home-grown wheat products. Taking into account the high bioavailability of Cd in rice grain, the limit for Cd in wheat should be higher than the limit for rice. 5.4.1. Many factors affect Cd accumulation by crops Identification of limits for Cd in fertilizers is remarkably complex because so many different aspects of agriculture can influence Cd accumulation and bioavailability. It is generally agreed that the first adverse effect of excessive lifetime Cd ingestion is renal tubular dysfunction of humans. Absorption of crop Cd is strongly affected by human nutritional status, the bioavailability of crop Cd (especially Fe, Zn, and Ca status of the exposed individual and the diet), and the presence of fiber and phytate in the diet. Humans may

Food Safety Issues for Mineral and Organic Fertilizers

67

ingest crops which contain high levels of Cd from production on Cdenriched soils, and the potential for risk is related strongly to the years of high Cd crop ingestion, and the fraction of diet grown on high Cd soils. Tobacco is another very significant source of human kidney Cd with smoking of normal cigarettes doubling kidney Cd at age 50 compared to non-smokers (Elinder et al., 1976). Natural soils vary widely in total Cd, from below 0.1 to over 2 mg kg
1 in normal agricultural soils, and over 100 mg kg
1 in rare highly Cd mineralized soils. Soils may be contaminated from very different sources of Cd including: aerosol emissions of industrial sources; Cd in irrigation water contaminated by industrial sources; manure; biosolids; composts; fertilizers (especially P-fertilizers); and other soil amendments. The Cd:Zn ratio of the source can vary from very low (commercial purified Zn), through Zn ores, to P-fertilizers and Cd-only sources such as Cd-pigment and Cd-plastic stabilizer wastes. Cd in soils has very wide variation in phytoavailability due to soil, crop, and environmental/management factors. The crop and soil factors which most strongly influence crop concentrations of Cd are 1) crop species and cultivar (Grant et al., 2008); 2) soil pH; 3) soil and fertilizer Zn; 4) P-fertilizer rate and Cd concentration; 5) chloride in soil or irrigation water (McLaughlin et al., 1999); 6) use of chloride containing fertilizers; 7) rate and form of N-fertilizer; 8) soil drainage; 9) preceding legume or Cd accumulating crop species (Oliver et al., 1993); 10) tillage; and 11) Zn deficiency of crop/soil (Oliver et al., 1994). When so many factors can influence Cd accumulation in crops, and the bioavailability of Cd in crops, it becomes difficult to demonstrate a relationship between Cd concentration in a P-fertilizer and risk from that Cd (Grant and Sheppard, 2008; McLaughlin et al., 1996; Mench, 1998). Only a few papers have examined the complex multiple interactions possible between tillage, fertilizer placement, Cd concentration in the applied fertilizer, form and amount of N-fertilizer and multiple years (e.g., Gao et al., 2010), and such studies show the difficulty in interpretation of Cd phytoavailability research. The major factors (soil pH, NH4eN rate, chloride, crop, cultivar, P-rate, tillage, N-placement, etc.) can be observed in many studies, but the interactions can overwhelm the major factors when one looks at the details. All the interactions make development of recommendations for farmers quite difficult except to avoid or minimize Cd accumulation in cropland, especially rice and tobacco cropland. Another topic which should be kept in mind in evaluation of the existing literature is that Cd plant uptake results of greenhouse pot experiments are not closely related to results of the same plant and soil in the field. In a comparison of greenhouse pots, outdoor microcosms and field plots, deVries and Tiller (1978) evaluated Cd and other element uptake by lettuce and onion from a soil treated with a biosolids. Uptake was

68

Rufus L. Chaney

4-times higher in greenhouse grown lettuce than field grown. This is believed to result from restriction of roots to the treated soil in pots, but not in the field or microcosms, the higher fertilizer density in pots, and the water use patterns in a greenhouse vs. outdoors. Many flaws in pot studies have been noted in Cd research; perhaps the most important common error is application of Cd without Zn, and use of high added rates of Cd that would never be allowed in the environment. Study of Cd addition at rates which induce biosynthesis of phytochelatin have been shown to be irrelevant to environmental exposures to the usual Cd þ Zn combination in the real world (McKenna and Chaney, 1995). Many misunderstandings of the effect of soil factors, food-chain factors, and human factors in risk from soil or fertilizer Cd to humans cause controversy about fertilizer Cd limits. The following discussion attempts to describe these factors and clarify how foods are protected from excessive Cd in nearly all circumstances. For particular crops, fertilizers with lower Cd concentrations are clearly indicated (tobacco, subsistence rice). But for other crops, extreme low Cd:P2O5 limits are not supported by research on food-chain transfer of Cd from applied fertilizers. Soil pH is usually found to be a highly significant factor in crop Cd accumulation, and because of this effect, any changes in soil pH (especially in the rhizosphere) due to form and rate of N-fertilizer may significantly affect Cd levels in plants. Literally hundreds of papers have reported the effect of decreasing pH raising crop Cd, and the effect of especially NH4eN increasing crop Cd (e.g., Grant et al., 1999; McLaughlin et al., 1996; Chaney, 2010). An interesting field test of the effect of N-fertilizer form and rate, and the Cd concentration in an applied P-fertilizer was conducted with wheat in the field in Sweden (Wångstrand et al., 2007). As expected (NH4)2SO4 caused greater Cd uptake than NH4NO3, which caused greater Cd uptake than Ca(NO3)2 (which causes the rhizosphere pH to rise). But even increased rate of Ca(NO3)2 caused increase in wheat grain Cd concentration compared with no added N-fertilizer. That increase may be due to effects of N on root growth, or increasing the Cd sink in the grain, but it is seen in many evaluations of grain Cd and grain N. Soil redox is especially important in Cd accumulation in rice grain. From the first studies of rice and Cd, it became apparent that little Cd was accumulated in rice shoots until the flooded paddy was drained (e.g., Takijima and Katsumi, 1973). Late rains reduced grain Cd. Careful studies showed that while the soil was flooded, Cd solubility and phytoavailability was very low. This appears to result both from the rise in pH and the formation of CdS in the flooded soil (Khaokaew et al., 2011). In recent research to characterize use of rice for phytoextraction of soil Cd, Murakami et al. (2009) reported that before draining the paddy in July, shoots contained less than 2 mg Cd kg
1 DW, while at

Food Safety Issues for Mineral and Organic Fertilizers

69

harvest, the shoots contained 70 mg Cd kg
1 DW. Further, Simmons et al. (2008) showed that in order to predict Cd uptake by rice from extractable Cd in soil, it was important to sample and immediately analyze the moist field soil so that the soil redox and pH during grain filling would be compared to grain Cd. And although grain and shoot Cd can rise by more than 10-fold after draining the field, and leaf Zn can be increased, grain Zn is hardly changed after draining the field. The failure to bring Zn along with Cd into the grain is a key part of why rice can cause so much higher Cd risk than other crops which do not exclude Zn under soil conditions where Cd is accumulated. For example, in the Simmons et al. (2003) report on rice accumulation of Cd and Zn from mine waste contaminated paddies in Thailand, the brown rice grain grown in a high Cd and Zn-contaminated field (229 mg Cd and 6420 mg Zn kg
1, pH 7.9) contained up to 2.64 mg Cd kg
1, but only about 22.2 mg Zn kg
1; but Chinese cabbage grown in the same field in another season accumulated 5.8 mg Cd and 643 mg Zn kg
1 DW (Simmons, Angle and Chaney, unpublished). 5.4.2. Cd in phosphate fertilizers Nearly all P-fertilizers naturally contain higher Cd than soils. Some of the earliest data showing higher crop Cd on P-fertilized soils came from Australia where the unusual soil conditions and high Cd rock phosphate source used to produce superphosphate for Australia promoted Cd uptake by crops by the time it was identified (Williams and David, 1973). High Cd superphosphate produced from high Cd phosphate rocks from Christmas and Nauru Islands was applied across Australia for decades. Some Australian soils promote Cd uptake because of acidity, and others because of chloride (see below). Williams and David (1973) found significant Cd accumulation in tillage depth soils (well correlated with P accumulation in the soils), and higher Cd in crops grown on the fertilized soils. After other research was completed, it became evident that not all P-fertilizers or soils comprised equal risk. Over the decades, many wellconducted field experiments in various nations have evaluated the value of different phosphate fertilizer products. More recently, many of these sites were examined for plant uptake of Cd in relation to fertilizer P application rate and fertilizer Cd or Cd:P concentrations. In these studies, when the P-fertilizer contained low levels of Cd, no increase in accumulation of Cd in soils or plants were observed even after many years of cropping (Mortvedt, 1984, 1987; Mortvedt et al., 1981; Rothbaum et al., 1979; Mishima et al., 2004; Richards et al., 1988; Isermann, 1982; Sauerbeck, 1982; Singh and Myhr, 1998; Smilde and van Luit, 1983), while when high Cd concentration phosphate fertilizers were used, soil Cd increased and crop Cd may have been increased over time, but still far below crop Cd levels needed to cause excessive human exposure over decades.

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In some other studies of this question, soil from highly fertilized fields was used in pot tests of Cd uptake. When soil accumulation of Cd was appreciable, plant uptake of Cd was strongly increased by the historic P and Cd application (Mulla et al., 1980). And when high Cd phosphates were applied to light-textured strongly acidic soils in pot studies, lettuce or other crops accumulated increased amounts of Cd compared to very low Cd P-fertilizers (Reuss et al., 1978; Singh, 1990). In other field experiments of short duration, even when the applied phosphate source had relatively high Cd concentration, the applied P and Cd in the different products (63 vs. 379 mg Cd (kg P)
1 on average) had little or no effect on Cd uptake by potato (McLaughlin et al., 1995); but the locations where the tests were conducted had large differences in Cd accumulation in potato tubers. The present application of phosphates with varied Cd levels showed low effect on Cd in several crops while NH4eN fertilizer rate more strongly affected Cd accumulation (Grant et al., 2010; Singh, 1990; Mortvedt et al., 1981). Cd accumulation was more related to the presence of high levels of chloride from irrigation water than soil Cd or even pH (McLaughlin et al., 1999). In these studies, application of KCl as the K fertilizer did not significantly increase Cd uptake into potato tubers compared to use of K2SO4, but perhaps the KCl was overwhelmed by the chloride in irrigation water. In another field test, using KCl caused a significant 30% increase in potato tuber Cd compared to K2SO4 (Sparrow et al., 1993, 1994). Perhaps because the high Cd:Zn ratio of P-fertilizers caused the Cd accumulated in Australian soils to have high Cd:Zn ratio, addition of high rates of Zn fertilizer (up to 100 kg Zn ha
1 tested) significantly reduced Cd in potatoes (McLaughlin et al., 1995). Other field studies have shown the effect of nitrogen rate and especially the NH4eN rate on Cd uptake by many crops (Grant and Sheppard, 2008; Grant et al., 1999; 2002; McLaughlin et al., 1996). N rate usually has a significant effect even when the Cd concentration in the applied Pfertilizer has no effect on plant Cd. Even increasing P rate caused an increase in plant uptake of Cd regardless of Cd level in the P-fertilizer (Grant et al., 2000). Further, preceding legume or Cd-accumulator crops (sunflower, flax, etc.) caused a significant increase in wheat Cd in several studies (Oliver et al., 1993; Grant et al., 2010). And tillage practices may also influence crop Cd concentrations (Oliver et al., 1993). In another study, the previous crop sunflower shoots were removed from the field which reduced Cd accumulation by wheat (Khoshgoftarmanesh and Chaney, 2007). These many interactions make it difficult to predict Cd accumulation in crops based on fertilizers or other agronomic factors despite the clear addition of Cd in some high Cd concentration P-fertilizers over time (McLaughlin et al., 2000).

Food Safety Issues for Mineral and Organic Fertilizers

71

While many assume that use of commercial phosphate fertilizers must cause higher Cd accumulation in crops than would organic farming methods, several field comparisons have been conducted. Jorhem and Slanina (2000) compared wheat in controlled field experiments and rye, carrots, and potatoes grown on commercial farms using either organic or conventional methods. There were no significant differences in Cd levels in rye, carrot, or potato over time, while wheat comparisons were inconsistent. They concluded that organic farming, at least in the short term, does not necessarily result in reduced levels of Cd in foods of vegetable origin. A similar outcome was reported by Harcz et al. (2007). 5.4.3. Cd in by-product zn fertilizers In the 1970s, some commercial P-fertilizers contained more than 400 mg Cd kg
1. Some fertilizer manufacturing by-products with more than 1000 mg Cd kg
1 were marketed in the US (CDFA, 1998). After a newspaper article in Seattle, WA, warned about the potential for contamination of fertilizers manufactured from hazardous industrial byproducts (Wilson, 2001), wide concern was raised. The product in question was manufactured from a hazardous fume waste of electric furnace steel production which contained substantial levels of Zn, Pb, Cd, As, and dioxins (US-EPA, 2002). A risk assessment conducted by US-EPA determined that limits were needed for such industrial byproduct derived fertilizer products, and they specified that the Zn must be separated from the byproduct to become equivalent to normal purified Zn (US-EPA, 2002). The regulation requires that Zn fertilizers manufactured from recycled hazardous materials (such as steel mill fume waste) contain no more than 1.4 mg Cd per 1% Zn in the product. For a typical ZnSO4,H2O product with 36.5% Zn, the Cd limit would be 50.7 mg kg
1; or for ZnSO4,7H2O with 22.7% Zn, the Cd limit would be 31.8 mg kg
1. The ratio of Cd:Zn thus has to be below 0.00014 which greatly limits Cd uptake by plants. Application of the unregulated products would have raised soil concentration of the contaminants. But because the Cd:Zn ratio was much lower than 1:100, excessive Cd accumulation in crops was not observed even with high cumulative application rates of these products (Mortvedt, 1985; Westfall et al., 2005). Zn in such Zn-fertilizer products inhibits uptake of Cd in the product. Increasing rates of product application over 5 repeated applications actually reduced Cd levels in lettuce shoots in a repeated application pot study by Kuo et al. (2004). The Zn product tested contained 469 mg Cd and 178,000 mg Zn kg
1, with Cd:Zn ratio of 0.00263 g g
1. But the application of such industrially-contaminated products on cropland was not acceptable due to the high levels of As, Pb and dioxins, and purification of the Zn from the recycled material was required if for no other reason than the dioxin present.

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Utilization of ground waste rubber has been tested as a Zn fertilizer. Tire rubber contains 1e2% Zn, and is generated in large quantities. Interestingly, rubber uses purified Zn such that the Zn in rubber is as low in Cd as required for byproduct Zn fertilizers by US-EPA (2002). Taheri et al. (2011) showed that ground rubber was an effective Zn fertilizer on Zn deficient soils and that adding ground rubber as a Zn fertilizer could reduce Cd in wheat grain. 5.4.4. Limits on Cd in phosphate products US States started to adopt limits for trace elements in fertilizer products, and the American Association of Plant Food Control Officials (AAPFCO) adopted recommended limits (Table 2). Other nations promulgated and the EU considered limits for Cd in P-fertilizer products with a schedule to reduce the allowable concentrations from 60 to 40 to 20 mg Cd (kg P2O5)
1 (equal to 133, 89, and 44 mg kg
1) product in triplesuperphosphate products containing 45 % P2O5 (EU, 1988) [see also the explanation of how these limits were developed in Cupit et al. (2002) and de Meeûs et al. (2002)]. But in 2011, the EU has not yet adopted regulations on Cd in Pfertilizers; EU expert groups reviewing the proposal suggested that the 20 mg Cd kg
1 P limit was more restrictive than needed to protect food safety (CSTEE, 2002). Some perspective on Cd and other trace elements in fertilizers is needed. First, historically, some fertilizer products have contained high levels of trace elements (Table 1) (Chen et al., 2007; Mortvedt, 1996; McLaughlin et al., 1996). A few fertilizers were contaminated by industrial sources, but some phosphate ores were naturally rich in Cd from their marine sources (Grant et al., 1999; McLaughlin et al., 1996, 1999; Nziguheba and Smolders, 2008) (Table 2). Others (e.g., Kola) were low or very low because of their ore source. Some were essentially devoid of Cd (electric furnace manufactured products) due to the production methods. But electric furnace phosphates are considerably more expensive than products made using sulfuric acid and they are no longer used in most commercial fertilizer products. So it is important to look at the evidence about potential risk from fertilizer Cd in human foods. As noted above, rice diets promoted Cd disease because Cd in rice had higher bioavailability than Cd in other foods. Further, although nearly all geogenic source Cd has high levels of accompanying Zn (Cd:Zn ratio lower than 0.010), rice grain was not increased in Zn when grown on the contaminated paddy soils (Fukushima et al., 1973; Simmons et al., 2003). Other grains and vegetables always have a large increase in Zn when Cd is increased by common Cd sources to cropland. P-fertilizers have less Zn per unit Cd than Zn ores (usual Cd:Zn ratio in phosphates is about 0.16 vs.<0.010 in Zn ores) and most sources of Cd contamination

73

Food Safety Issues for Mineral and Organic Fertilizers

except NieCd battery manufacturing wastes or Cd-pigment production or disposal wastes (Chaney, 2010). Another crop with unusually strong Cd accumulation is tobacco (Bell et al., 1988), and the Cd in tobacco enters the lungs of smokers and is well absorbed (Elinder et al., 1976). Tobacco is a stronger source of Cd in human kidney than is food excepting rice grown on contaminated soils. So Cd enrichment of rice and tobacco cropland require special attention. Tobacco grown on mine waste contaminated soils contributed significantly to the Cd health effects at one location in China (Cai et al., 1990), and in Thailand. Production of tobacco on Cd-enriched soils is contraindicated, and it has been recommended that P-fertilizers for use in tobacco production should be low in Cd (Lugon-Moulin et al., 2006). As noted above, many studies have been reported for Cd accumulation by crops grown on long-term phosphate fertilized soils. Because some locations received applications of Cd-rich P-fertilizer products for decades, some papers reported significant increase in crop Cd due to P-fertilizer applications (e.g., Williams and David, 1973). But in many other studies, even very long-term field experiments, no significant increase was found in crop Cd despite small increases in soil Cd (e.g., Mortvedt, 1987; Jung et al., 1979). In other cases, manure applications applied more Cd to field plots than did fertilizers and aerosol emissions, but the manure applications caused reduced levels of Cd in crops (Jones and Johnston, 1989). Looking at all common sources of Cd for cropland, Nicholson et al. (2003) found that deposition of aerosol metals exceed P-fertilizers and biosolids as Cd source in the UK soil budget (Table 3). And livestock manures ranged widely in Zn, Cu and Cd concentrations (Nicholson et al., 1999) (Table 4). In addition, different farm types use different soil amendments, hence giving different potential Cd accumulation. For example, farms which concentrate on animal production purchase off-farm feedstuffs and feed supplements and apply their manure locally (e.g., Bengtsson et al., 2003), Table 3 Zn, Cd and Cu accumulation and Cd:Zn of accumulated metals in soils of England and Wales from different sources Zn Sources

mg yr

Aerosol Deposition Livestock Manure Biosolids Phosphate Fertilizers

2457 1858 385 213

(Nicholson et al., 2003).

Cd

Cu

L1

Cd:Zn g gL1

21 4.2 1.6 10.0

631 643 271 30

0.0085 0.0022 0.0042 0.047

74

Table 4

Zn and Cd levels in livestock manures and Cd:Zn ratios of the mean Cd and Zn concentrations Zn Mean

Species

Dairy Cattle FYM Dairy Cattle Slurry Beef Cattle FYM Beef Cattle Slurry Pig FYM Pig Slurry Broiler/Turkey litter Layer Manure

Cd Range

Mean

Range

mg kgL1 dry matter

153 209 81 133 431 575 378 459

99e238 <5e727 41e274 68e235 206e716 <5e2500 208e473 350e632

0.38 0.33 0.13 0.26 0.37 0.30 0.42 1.06

<0.10e0.53 <0.10e1.74 <0.10e0.24 0.11e0.53 0.19e0.53 <0.10e0.84 0.20e1.16 0.44e2.04

Cd:Zn

Cu

Mean

Mean

g gL1

mg kgL1 dry matter

0.0025 0.0016 0.0016 0.0020 0.00086 0.00052 0.0011 0.0023

37.5 62.3 16.4 33.2 374 351 96.8 64.8

Range

26e56 <1e352 10e28 18e49 160e780 <1e807 46e173 49e75

(Nicholson et al. 1999). Rufus L. Chaney

75

Food Safety Issues for Mineral and Organic Fertilizers

and do not purchase P-fertilizers or use biosolids because they don't need P and seek to avoid excessive P accumulation in their soils. Keller and Schulin (2003) took into account different amendment applications on qualitatively different farms to examine the Zn and Cd balance. As shown in Table 5, livestock farms had the highest rate of Cd accumulation of all farm classes. Further, the limited amounts of total biosolids in a jurisdiction means that few farms will be able to apply biosolids as a P and N fertilizer, strongly limiting the Cd accumulation from modern biosolids (now much lower in Cd than in the 1970s). Although biosolids with as high as 3700 mg Cd kg
1 DW were reported in the older literature (Sommers, 1977), industrial pretreatment of wastewaters, and reduction in Cd use in products has greatly reduced Cd in biosolids. For example, in a study of Pennsylvania biosolids metals levels over a 22-year period, Stehouwer et al. (2000) showed the decline from median Cd of 6 mg kg
1 to 1.5 mg kg
2 such that the 90th percentile of biosolids Cd was only 4 mg kg
1. And in the 2009 Targeted National Sewage Sludge Survey (US-EPA, 2009), the geometric mean Cd was 1.97 mg kg
1 (full range, 0.21 to 11.8 mg kg
1), while the mean Cd:Zn was 0.0028 [range 0.0006 to 0.0106 g Cd (g Zn)
1]. Thus, except in subsistence rice and tobacco production (Bell et al., 1988), modern biosolids do not cause Cd risk to humans or the environment because of their low Cd:Zn ratio and high Cd adsorption strength (Kukier et al., 2010).

Table 5 Estimated Cd accumulation in soils of Sundgau Canton from different sources, and for different farm types Arable Farming

Dairy& Mixed

Animal Husbandary

Whole Region

g haL1 yrL1

Input Fluxes Manure Biosolids Commercial Fert. Deposition Output Fluxes Crop Removal Leaching Net Flux: (Keller and Schulin, 2003).

<0.1 0.4 0.6

0.7 0.1 0.2

8.2 <0.1 <0.1

0.6 0.1 0.4

2.1

2.1

2.1

2.1

0.6

1.9

0.6

1.4

0.4 2.3

0.4 0.8

0.4 9.1

0.4 1.4

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Rufus L. Chaney

5.4.5. Effect of chloride on cadmium accumulation in crops One of the most remarkable aspects of crop Cd accumulation and foodchain Cd risk is the role of soil chloride in Cd uptake. This effect was first demonstrated by Bingham et al. (1984) in pot studies. Then McLaughlin et al. (1994) found that Cd in potato tubers was more strongly affected by soil chloride levels than other soil factors they evaluated. This was soon confirmed for Cd in sunflower kernels (Li et al., 1995), bread wheat (Weggler-Beaton et al., 2000), and durum wheat (Norvell et al., 2000). McLaughlin and cooperators conducted many studies to better understand this interaction between Cd and chloride. For example, increasing chloride in an unbuffered nutrient solution causes some of the Cd2þ to be converted to CdClþ, lowering the activity of Cd2þ, but increasing uptake of Cd by Swiss chard (Beta vulgaris L. var. cicla) (Smolders and McLaughlin, 1996a). Chard was used in these tests because it is highly tolerant of NaCl compared to many other crops and strongly accumulates phytoavailable Cd. Finally, using a chelating resin-buffered nutrient solution approach which maintained constant activity of free Cd2þ in solution, both dissolved total Cd and Cd uptake by Swiss chard was significantly increased when chloride was added to the solution (Smolders and McLaughlin, 1996b). The conclusion of these studies was that when chloride was present at high levels, Cd2þ activity did not control Cd uptake as previously hypothesized. And with the resinbuffered system, it is clear that the CdClþ or CdCl02 species must be taken up by roots. Another alternative was that the formation of chloride complexes increases the pool of diffusible Cd in the soil up to the root membrane, and because the chloride complexes can rapidly exchange free Cd2þ, they may maintain higher levels of free Cd2þ at the root cell membrane (Degryse et al., 2006a, 2006b). The pronounced effect of field chloride levels were also identified in fields of durum wheat where soil chloride varied across the landscape (Norvell et al., 2000; Wu et al., 2002). Interestingly, increased soil solution sulfate also increases dissolved Cd as the sulfate-complex, but that has little effect on Cd uptake by plants (Bingham et al., 1986; McLaughlin et al., 1998). The formation constant for the sulfate-complex is similar to that of chloride. In the Li et al. (1995) study with sunflower kernels, the same highly calcareous soil series occurred at different locations with different chloride levels but both had high soil solution sulfate. But kernel Cd was more than doubled in fields with higher soil chloride with no effect on yield. And among fields across Australia, Cd in potato was more affected by chloride than by soil total Cd or soil pH, illustrating the overwhelming potential effect of soil chloride. In predicting potato Cd, both chloride in the soil at planting and the irrigation water supply were

Food Safety Issues for Mineral and Organic Fertilizers

77

important (McLaughlin et al., 1999). Chloride accumulates in poorlydrained soils from KCl fertilizers and other inputs, and can be considerably increased by geogenic sources (Wu et al., 2002). Particularly in alkaline soils, chloride increases Cd uptake more than other soil factors studied, while in acidic soils chloride is less important apparently because Cd solubility is so much greater that chloride complexation is less significant (Hattori et al., 2006). Taken together, these findings indicate that the CdClþ species must be absorbed by roots, but the specific transporter remains unknown; it is considered unlikely that the Zn transporter plays a role in uptake of Cd from CdClþ such that the overall rate of Cd uptake is increased. Most publications on the effect of chloride on Cd uptake support the view that chloride causes higher dissolved Cd due to formation of chloride complexes (but not more than sulfate), and that the rapid exchange of chloride at the root cell membrane lowers charge and promotes uptake. Clearly diffusion is increased, but the difference between the effects of chloride and sulfate is not fully explained by this model. McLaughlin et al. (1998) suggest that the (CdSO4)0 complex is taken up equivalent to Cd2þ because increasing complexation of Cd2þ by sulfate did not reduce Cd uptake by chard. Alternatively, Khoshgoftar et al. (2004) reported that increased soil chloride reduced the activity of free Zn2þ at the root, increasing Zn deficiency stress on the plant and promoting Cd uptake by the root Zn transporter. This model for Zn deficiency stress increasing Cd uptake was also found in studies of soil amended with a Cd-contaminated biosolids with low Zn. Liming the soil actually induced Zn deficiency in lettuce and strongly increased Cd uptake by lettuce. Zn fertilization to prevent liming-induced Zn deficiency reduced lettuce Cd from 19.3 to 1.95 mg Cd kg
1 DW (Chaney et al., 2006). The increase in crop Cd upon liming is in strong contrast with many studies of the effect of pH on Cd uptake (Chaney, 2010). The best available data show that both Cd and Zn are transported on the root Zn2þ transporter (Hart et al., 2002; 2005). Despite an earlier paper suggesting that the root ferrous transporter (IRT1) transported Cd, Cohen et al. (2004) later clarified that if the activity of Cd2þ and other ions in soil solution are taken into account, the ferrous transporter (IRT1) is not involved in Cd2þ uptake. 5.4.6. Retention of Cd from feedstuffs in tissues of livestock Most crops are consumed by livestock, and little of the ingested Cd remains in the livestock tissues consumed by humans. The EU Cd Risk Assessment (EFSA, 2009) ignores the nearly total exclusion of Cd from the carcass of livestock even when their diets are enriched in Cd. This has been tested in experiments where biosolids were added to cattle diets to simulate the

78

Rufus L. Chaney

inadvertent ingestion of biosolids during grazing (Kienholz et al., 1979), or biosolids were used to produce corn silage fed to goats and sheep. In the Kienholz et al. (1979) study, feedlot cattle were fed diets with 0, 4, or 12% biosolids (dry weight basis) for 94 days before slaughter and analysis of tissues. The biosolids contained 21 mg Cd and 1500 mg Zn kg
1 DW, higher than modern biosolids after pretreatment has become widely enforced. They measured that <0.04% of diet Cd was retained in the whole carcass of the cattle at the end of the experiment. Table 6 shows the balance of Cd in a three-year experiment with dairy goats reported by Bray et al. (1985). Even with the highest rate of biosolids, and the high Cd concentration and Cd:Zn ratio in the biosolids applied, the goats only retained 1.9 mg Cd in liverþkidney of the 3980 mg of Cd which were in their feed. Hinesly et al. (1985) similarly found a very low fraction of the Cd in corn and soybeans fed to chickens to be retained in the chicken bodies, and none in eggs even for high Cd biosolids. Decker et al. (1980) found no increase in tissue Cd in cattle which grazed on low Cd biosolids amended pastures or were fed low Cd biosolids in diets. Similarly, adding 5 ppm Cd as salt to diets of diary cows did not cause any increase in milk or meat tissues Cd concentration (Smith et al., 1991). Some earlier studies which used addition of Cd salts to livestock diets found accumulation of Cd in tissues, but when Zn is added along with the Cd, much lower accumulation is observed (Stuczynski et al., 2007). In the latter study, forage from uncontaminated soil and from a remediated Zn smelter slag were fed to calves, compared with additions of Cd or CdþZn to the uncontaminated soil grown forage to match levels in the contaminated soil grown forage. The increased Cd in the forage grown on the remediated site caused a small significant increase in kidney Cd (0.53 vs. 0.17 mg kg
1 wet weight from control forage), while Table 6 Accumulation of Cd by goats fed corn silage grown on biosolidsa amended soils for three years; silage composition for year-3 Corn Silage

Feed

Liver

Kidney

Muscle

LDK

Cd

Cd

Cd

Cd

Cd

Cd

mg/3 yr

mg/kg DW

Zn

Treatment mg/kg DW

Control Low Medium High a

<0.06 1.39 2.73 5.26

34.5 73.5 74.6 107

<40 1290 2180 3980

0.26 1.72 2.10 2.94

3.09 10.9 24.7 22.4

mg

<0.06 <0.05 <0.07 <0.12

0.17 0.77 1.32 1.41

Biosolids contained 1700 mg Zn kg-1 and 105e186 mg Cd kg-1, with high Cd:Zn. The high Cd and Cd:Zn ratio allowed higher Cd in crops and goat tissues than commonly found for high quality biosolids (Cd source was a Cd-Ni battery factory). (Bray et al., 1985).

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the Cd salt addition caused a large significant increase (2.10 mg kg
1) which was reduced considerably by also adding the normal Zn which accompanies Cd in such forages grown on contaminated sites (0.78 mg kg
1). Thus the normal livestock production feed-chain comprises very strong limitation of Cd transfer from soils and soil amendments to humans. 5.4.7. Exceedance of kidney Cd limits in sheep due to historic P-fertilizer Cd Because the P-fertilizers used in Australia and New Zealand for decades were rich in Cd, soils, forages, and sheep kidney and liver of older animals showed significantly higher Cd accumulation than where low Cd fertilizers were used (Roberts et al., 1994; Loganathan et al., 1999). Thus, Australia and New Zealand have set low limits on Cd in kidney and liver of sheep in order to meet import limits in the EU. Factors which cause kidney of older sheep to exceed these limits may clarify if similar outcomes would be expected in other countries. A number of factors interact to create a remarkable Cd transfer case in some Australian and New Zealand sheep. First, the long application of high Cd phosphate fertilizers from Nauru rock phosphate, soil acidity or chloride, and age of the animals are each important. Only kidneys of older animals fail to comply with EU Cd limits. After detailed investigation, it was also discovered that unusual accumulation of Cd but not Zn by capeweed (Arctotheca calendula) [an introduced weedy legume which infests pastures and grows well in winter] causes the total Cd and the Cd:Zn (already high from the high Cd:Zn ratio of phosphate fertilizer) of the sheep diet to be substantially increased compared to animals grazing fields where capeweed was controlled (McLaughlin et al., 1996, 1997; Hamon et al., 1997; Bramley and Barrow, 1994). Also, liver and kidney limits have been based on assumptions of regular lifetime consumption of liver and kidney rich in Cd. This scenario is unrealistic and generates unnecessarily low Cd limits for kidney and liver, and in fertilizers and other soil amendments. 5.4.8. Has there been an important increase in crop Cd over time? Concern about increasing food Cd over time was initiated by a report of Kjellström et al. (1975) who analyzed wheat samples from different harvest years from a location in Sweden. And increased by the assumption that Cd in applied fertilizers necessarily causes increased Cd in crops. The samples they analyzed indicated that crop Cd had doubled in 50 years, so they expressed concern about this pattern. Andersson and Bingefors (1985) tested more samples from this farm operated by a plant breeding company and found large annual variation in grain Cd over time. The wheat was grown on different fields in different years due to normal crop rotation; different fields had different pH levels and Cd

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levels from parent materials, which caused large year-to-year variation in crop Cd concentration. Thus when field-to-field variation and cultivar differences were considered, it became unclear whether a meaningful increase in wheat grain Cd had occurred. There was no increase in barley grain Cd. A more recent comparison of Cd in wheat grown on long-term soil fertility field test plots from 1967 through 2003 showed a significant decline in grain Cd, attributed to reduced aerosol deposition of Cd in Sweden (Kirchmann et al., 2009). Lorenz et al. (1986) conducted a similar evaluation of changes in crop Cd from long-term samples maintained at a location in Germany, but found no increase in wheat Cd over time. Recently Jorhem et al. (2001) tested flour from the period 1983e1997 and also found no significant increase in wheat flour Cd over time. Rothamsted long-term field plots tested the change over time of soil and crop Cd, finding increased grain Cd on very acidic soils (but not if the soils were limed) but decreased grain Cd with considerably higher amounts of manure-applied Cd ( Jones and Johnston, 1989). Further, a long-term comparison of Cd in foods in Finland also found significant decline over time rather than increase (Ekholm et al., 2007). Examination of the balance of Cd, Zn, P, and some other elements in cropping systems has been undertaken by scientists in several countries. By examining the composition of all sources of Cd for soils (fertilizer, irrigation water, aerosol deposition, manure, biosolids, etc.), and estimating the normal rate of nutrient amendment application limited to avoid excessive application of N or P to soils, they could evaluate the role of farming system and soil amendments on long-term Cd balance. The results are quite interesting because they are counterintuitive. With the composition of P-fertilizers being used in 2000, both aerosol deposition and manure applications applied more Cd to farm types of Switzerland (Table 5) than did fertilizers (Keller and Schulin, 2003). Livestock production farms import feed supplements and apply the livestock manure onto their fields, but need no additional P-fertilizer and may need no N-fertilizer. And would apply no biosolids because that would add additional unnecessary P. A similar conclusion was reached in the assessment of long-term accumulation of Cd in soils and crops of Canada (Sheppard et al., 2009). 5.4.9. Geogenic Cd enrichment of cropland Cadmium, arsenic and other potentially problem elements are part of the natural world. Although soil Cd is accompanied by 200e500-fold higher Zn in most soils, a few soils developed from mineralized rocks such as marine shale (Lund et al., 1981) or alum-shale (Mellum et al., 1998) (also called iron sulfate soils) which are Cd mineralized without usual Zn co-mineralization. Soil development in Somerset, UK, included some Zn ore materials which caused higher Cd in wheat grain (Chaudri et al., 1995). Natural

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mineralization of marine shale parent rocks on the western side of Salinas Valley, CA, has provided much Cd but little Zn to residual soils, and Cdenriched regional alluvial soils (see Holmgren et al., 1993; Chen et al., 2009). Combined with high chloride in irrigation waters, these soils promoted Cd accumulation by leafy vegetables (Burau, 1983; Wolnik et al., 1983; Chen et al., 2009). Alternative land use kept most of the high Cd residual marine shale soil from producing vegetables for decades, but changing economics changed farm practices. Interestingly, liming these soils did not reduce Cd accumulation by spinach (Burau, 1983), in strong contrast with the usual strong reduction in Cd uptake by crops when soil pH is raised by 1 or more units (Chaney, 2010). Chaney et al. (2009) tested the hypothesis that the low level of soil Zn, coupled with high Cd:Zn ratio, caused plants to experience functional Zn deficiency on such soils, causing plants to attempt to obtain more soil Zn but absorbing the Cd present as found with a soil very high in Cd from a contaminated biosolids (Chaney et al., 2006). When adequate Zn was applied along with the limestone to make soils with 9 mg Cd kg
1 calcareous, lettuce Cd fell from over 12 mg kg
1 DW to well under the 4 mg Cd kg
1 DW of the proposed Codex standard for leafy vegetables (Chaney et al., 2009). Perhaps the most remarkable example of geogenic Cd enrichment of soils was identified in Jamaica (Lalor et al., 1998). In some locations, natural soil Cd exceeds soil Zn! Geological research found that marine fossils provided the Cd to the parent rocks millions of years ago that formed these soils (Garrett et al., 2008). Because most of the soils are near neutral pH, very rich in Fe and Mn oxides, and not high in chloride, local crops do not contain extreme levels of Cd, although clearly above background levels (Howe et al., 2005). In contrast with the cases in rice Cd in Asia, clear evidence of renal tubular dysfunction has not been observed in Jamaica (Lalor et al., 2004; Lalor, 2008; Wright et al., 2010). With this example, other areas with similar soil genesis from marine shale sources need to be examined for potential food-chain Cd risk from mineralized soils. 5.4.10. Deliberate contamination of a fertilizer product Although most fertilizer manufacturers conduct careful quality control, a few have violated business norms. One egregious case occurred with a ZnSO4 product marketed by one or more Chinese chemical companies. The situation was discovered when importation of the ZnSO4 was sought in British Columbia, Washington, Oregon and California in 1999. Because of the Seattle newspaper articles (Wilson, 2001), each of these jurisdictions had fertilizer laws and enforcement agencies, so samples of the product had to be locally analyzed and the product registered on that basis. But the product had remarkable levels of Cd (Table 7), far above natural abundance (Chaney, 2010). One sample had as much Cd as Zn,

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Table 7 Cd and Zn concentration and Cd:Zn ratio of in samples of Cd-contaminated Zn fertilizer product delivered to northwestern US and Canada Cd

Zn L1

g gL1

DW

Sample

mg kg

China-Zn-1 China-Zn-2 China-Zn-4 China-Zn-5 CenesaZnSO4 Blue-Minb

46,400 72,800 215,000 199,000 7.1 49

Cd:Zn

345,000 313,000 216,000 230,000 320,000 420,000

0.135 0.233 0.995 0.865 0.000022 0.000127

a

Zn fertilizer grade product. Byproduct Zn fertilizer. (Washington State Department of Agriculture, unpublished data, 1999).

b

indicating that waste Cd was collected from other sources to mix with and adulterate the ZnSO4 ingredient. Subsequently, highly Cd-contaminated ZnSO4 products were sold in other countries. News reports of Cd contamination in South Africa, Kenya, Australia and the EU (France, Belgium) showed that the contaminated ZnSO4 exports continued in the 2000s (ec.europa.eu/food/ committees/regulatory/scfcah/animalnutrition/summary38_en.pdf). In these cases, the vendor certified the product contained low levels of Cd (e.g., <15 mg kg
1), but eventual analysis showed much higher Cd levels. Zn product used in animal feeds harmed livestock from the Cd present. And high Cd in the Zn fertilizer or livestock manure applied to crops caused crops to exceed import limits in the EU. Few details of these cases have been published, but the harm is persistent. A single application of this highly contaminated Zn fertilizer may include 2e5 kg Cd ha
1 when applying 5e10 kg Zn ha
1. This case is instructive. Simply having a good law on trace elements in fertilizers is not enough. These products harmed farmers in Australia and the EU which had clear fertilizer Cd limits. Enforcement of the law such as occurs in Washington State, California and Canada is required for adequate protection to be obtained (see Washington State Department of Agriculture web site). Analysis of commercial fertilizer product samples is required before registration and approval for sale; companies must confirm that their product remains below state limits in trace elements in reports to the WA Department of Agriculture. It is likely that highly Cd-contaminated Zn fertilizer products have been sold in many other countries and parts of China; but without monitoring, no data are available to estimate the extent of any problem. The People's Republic of China (PRC) is reported to have responded to complaints from the EU by agreeing to

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conduct government testing of any Zn fertilizer or feed additive products to be exported to the EU, and a PRC government analysis certificate would accompany any product sold in the EU. That does not protect other nations with fewer regulatory protections. A deliberately Cd-contaminated Zn fertilizer product is especially worrisome in Asia where use of such Zn fertilizer on rice soils (which often receive Zn fertilizer) could eventually cause human Cd disease to be discovered decades later in subsistence rice farm families.

5.5. Lead Because Pb is so strongly adsorbed or precipitated in soil and plant roots, plants do not accumulate high Pb concentrations even when soils are appreciably contaminated with Pb. Pb phytotoxicity does not occur in the general environment (Koeppe, 1981) even when soil Pb levels comprise risk to children. Urban soils are remarkably contaminated by historic automotive emissions, stack emissions, and house paint residues, particularly in the inner city (Mielke et al., 1983). Because Pb uptake into edible parts of plants is so small, soil ingestion is the most limiting risk from soil Pb, and high Pb soils are not recommended for gardens to limit soil ingestion exposure. But urban citizens commonly grow garden crops in such high Pb gardens around the world. Fortunately, application of phosphate fertilizers and compost products can strongly reduce the bioavailability of soil Pb to persons who ingest soil (Sterrett et al., 1996; Zia et al., 2011). A field test was conducted at Joplin, MO, near a Pb smelter where contaminated lawn soil containing about 3000 mg Pb and Zn kg
1 was treated with different soil amendments to attempt to reduce soil Pb risks. The study included feeding untreated and P-treated soils to human volunteers. Soil Pb bioavailability was reduced 69% by treatment with phosphoric acid, nearly as much by treatment with triplesuperphosphate, but less with rock phosphate (Ryan et al., 2004). Soil Pb was converted to chloro-pyromorphite, a form of Pb with very low bioavailability (Scheckel and Ryan, 2004). When high applications of phosphate are to be used to reduce soil Pb risk, products low in Cd should be selected to avoid the possible high Cd loading with a remediation application of phosphate. Lead arsenate pesticide was major source of Pb contamination of orchard soils (Peryea and Creger, 1994; Merwin et al., 1994). Numerous studies of Pb uptake by garden crops found little uptake (e.g., Elfving et al., 1978). But the US Food and Drug Administration found two samples of commercial carrots with anomalous levels of Pb, and advised the industry to find a way to prevent this from occurring again. The industry contracts for most carrot production, so they required growers to prove that their soils were not high in Pb, much like carrot growers

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have long been required to show their soils are not high in DDT. Codling et al. (2012) tested why carrots contained high Pb levels when grown in old orchard soils and found that the Pb was not soil particles adhering to the carrot peel layer, or Pb ions accumulated into the peel layer cells, but was within the peeled carrot. Chaney et al. (2010a,b,c) examined the localization of Pb in these carrots using m-X-ray fluorescence and found that the Pb was within the xylem. The xylem of root crops (which are actually expanded hypocotyls) grows through the edible root, so part of the absorbed Pb moving to the shoots is trapped in the xylem. In contrast, potato tubers are phloem fed and contain no increase in Pb when grown on the same Pb-rich orchard soils (Codling et al., 2012).

6. Trace Elements in Other Common Soil Amendments 6.1. Manure All livestock manure contains Cd, Cu, Zn, etc. from the feed ingredients and P-supplements. Nicholson et al. (1999) summarized the Cd, Zn, Cu, As and other element concentrations in manure from different livestock classes in the UK. The means and range for Zn, Cd, Cu, and the Cd:Zn ratio of the means are shown in Table 3. These data show that swine and poultry feeds are commonly supplemented with high levels of Zn and Cu, and may contain an As compound as a growth stimulant or antibiotic. Because of the high levels of Zn and Cu in swine manure, repeated manure applications are the dominant source of increase in soil Zn and Cu in many nations (Sheppard et al., 2009; Keller and Schulin, 2003). Accumulation of As in soils amended with poultry litter has been observed in some experiments (Adeli et al., 2007), while not in others (Codling et al., 2008). Because manure contains high phosphate levels, and phosphate competes more strongly than arsenate for adsorption sites in soils, the co-applied phosphate can cause leaching of applied As. So for the low Fe, light-textured Coastal Plain soils in the Codling et al. (2008) experiment, As was not increased after decades of poultry litter application. But in the Adeli et al. (2007) study, and other research with heavy textured and/or high Fe soils, poultry litter applied As remains in the plow layer depth (as do As from orchard sprays, herbicides and defoliants) (Staed et al., 2009). Despite low levels of Cu in normal forages used as ruminant feedstuffs, McBride and Spiers (2001) found high Cu in dairy manure. The source was found to be CuSO4 use in hoof baths to prevent a hoof disease in dairy cattle. The CuSO4 bath becomes fouled with manure and soil over

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time, and was commonly dumped into the liquid manure handling system. Dairy manure with 1000 to even 3000 mg Cu kg
1 DW have been reported. Tests have shown that livestock producers do not need to use the Cu baths continuously; and the fouled baths can be disposed with other solid wastes rather than dumped into the manure. Producers are presently advised to minimize manure Cu to prevent future problems with Cu accumulation in soils. Adverse effects of manure Zn and Cu have seldom been reported perhaps because the manure adds metal sorbent to amended soils, and normal near neutral pH management minimizes the potential for Zn or Cu phytotoxicity. All phytotoxicity of soil Zn, Cu and Ni occur at strongly acidic pH until extreme amounts of metals are applied, far more than in present manures (Chaney, 1993, 2010). Historically excessive application of Cu pesticides and fertilizers in Florida, USA, caused accumulation of Cu in sandy soils low in organic matter, and with low pH. These conditions promoted Cu phytotoxicity to citrus and vegetables (Reuther and Smith, 1954), which was counteracted by liming and adding FeEDTA to correct Fe deficiency induced by the Cu. In European long-term vineyards, applications of Cu pesticides caused accumulation of Cu to high levels such that some grasses suffered Cu toxicity (Cu-induced Fe deficiency) even at neutral pH which should ordinarily prevent Cu phytotoxicity. In this case, the phytosiderophores secreted by young roots of durum wheat were filled by Cu rather than Fe3þ, thus inhibiting the normal Fe uptake mechanism of grasses (Michaud et al., 2007). Similar effects on dicots are not observed in these soils, but the chlorosis of durum wheat is an adverse effect of historic Cu pesticides in vineyards and some excessively Cu fertilized soils. Concern about As in soils and land-applied poultry litter are discussed below. The tolerance of increased soil As is very complex, and the benefit of arsenicals in poultry diets have been demonstrated ( Jones, 2007). False claims about As retention in chicken carcasses (Lasky et al., 2004) have confused many about transfer of As to humans in the poultry meat consumed (Tsuji et al., 2007). But objective measurements have shown that poultry fed arsenicals have little or no increase in tissue As except in liver, and that is depleted during the required withdrawal period. So the most direct justification to stop use of arsenicals may lack a technical basis. As is described below, extreme claims about cancer risk from exposure to even below-background levels of As in soils, water and food suggest that even background levels of soil As cannot be tolerated. It seems unlikely for this to occur in nature.

6.2. Limestone Byproducts containing limestone equivalent have often been recovered during Zn and Pb ore beneficiation, and historically these materials were

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used as agricultural limestone. Some other byproduct limestones are quite low in trace elements (sugarbeet lime; acetylene lime). Chaney et al. (1988) reported that dolomitic limestone sold in the vicinity of Palmerton, PA, had usually been the “Horsehead” brand, a product of Zn ore beneficiation at the Friedensville, PA, mine of the New Jersey Zinc Company. Three samples contained 3.6, 3.9 and 8.4 mg Cd kg
1 dry weight, and 2500, 2810 and 4580 mg Zn kg
1 (Cd:Zn ratios are 0.00156, 0.00139 and 0.00173, all far below the ratios of ores and fertilizers). Normal agricultural limestone contains <0.3 mg Cd kg
1 and 21 mg Zn kg
1. Other Information indicated that a similar Cd and Zncontaminated mining by-product limestone from Tennessee was sold in several Southeastern states. Pb, Zn and Cd rich limestone byproducts were often marketed in Missouri, USA, and other states from the mines in the “tri-state area mining district” (KS, OK and MO). Several experiments were conducted to test the effect of using these products on trace element accumulation by garden crops (Davies et al., 1993; Wixson et al., 1984). As might be expected, use of limestone raised soil pH and reduced the phytoavailability of soil Zn and Cd very strongly. But application of a limestone product with 10 mg Cd kg
1 at five t ha
1 every five years would add 1.0 kg Cd ha
1 every century, more than estimated for application of P-fertilizers.

6.3. Steel production fume waste In a case in Tifton, Georgia, USA, a steel mill fume waste rich in Zn and limestone equivalent was sold as a substitute limestone source. The combination of light-textured soils and production of peanuts led to severe Zn phytotoxicity in peanuts grown on these soils several years later as the soil pH naturally fell due to rainfall and fertilizer use (Davis et al., 1995). Peanut is unusually sensitive to excessive Zn in acidic soils (Keisling et al., 1977). Fortunately, high Zn in the soil amendment restricts Cd accumulation in a peanut (data not reported for the Georgia case) as shown by McLaughlin et al. (2000).

6.4. Gypsum Gypsum (CaSO4$2H2O)is applied to soils as a Ca or sulfate fertilizer, as an soil aggregation aid to water infiltration, to prevent crusting, and to reduce erosion; to aid in alleviation of subsoil Al phytotoxicity, and as a Ca source to remediate sodic soils (contain excessive Na). In the past, mined gypsum and phosphogypsum (from the phosphate fertilizer manufacturing industry) were used in many nations. Mined gypsum does not commonly contain trace elements at levels which comprise concern. Phosphogypsum may contain Cd and Zn, and Ra and other isotopes which could limit use of

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this form of gypsum on land (Mays and Mortvedt, 1986; Rutherford et al., 1994). Phosphogypsum contains a variable fraction of Cd from the phosphate rock depending on the manufacturing process (see Wakefield, 1980). More recently gypsum manufactured during removal of SO2 from the exhaust of coal-fired power plants (Fluidized Gas Desulfurization Gypsum, FGDG) has been evaluated for beneficial use in agriculture. Until the 2000s, this gas treatment was conducted with fly ash removal and some of those products contained potentially phytotoxic levels of boron, and could have high levels of Se, As, etc. (e.g., Kost et al., 2005; Ransome and Dowdy, 1987). But when it became evident that the trace element contaminated FGDG would not be acceptable in agriculture, the industry tested changing the gas treatment train to remove fly ash first, then to remove the SO2 and produce clean FGDG. Power plants are required to remove the SO2, and by 2020 are expected to generate on the order of 30 million t yr
1 in the US. Fortunately this change to much cleaner FGDG has prevented unusual levels of nearly all trace elements from entering the FGDG. Only Hg, Se and As were somewhat higher than background soils, while Cd, Zn, and other trace elements are similar to soils and mined gypsum (Chaney et al., 2010). In the presence of the huge 2amount of SO24 , SeO4 uptake by crops is not increased, so the slight increase in Se in the FGDG does not comprise risk. Sulfate poisoning of cattle was observed when appreciable levels of gypsum were added to a feed supplement for range cattle (Raisbeck, 1982). Excessive sulfate intake by ruminants can cause a severe disease, polioencephalomalacia (Gould, 1998). Adherence of powdered gypsum on leaves following surface application on pastures could allow increased intake by grazing livestock, as could direct access to FGDG piles in a field. Thus, a waiting period for rainfall or forage growth should occur after surface application of powdered gypsum before ruminants may graze amended pastures, and ruminants should be fenced out of FGDG storage piles.

7. Long-Term Reactions of Trace Elements in Soils It has long been known that the Zn in ZnSO4 fertilizers added to deficient soils was slowly transformed to forms with lower phytoavailability (e.g., Boawn, 1976). A normal fertilizer application on a deficient soil may require additional Zn fertilization after 5 or more years. The mechanism of Zn becoming less phytoavailable over time has recently become better understood. Initially, it was thought that the added Zn became more strongly bound on Fe and Mn oxides, or diffused into micropores in the

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oxides where it could not reach soil solution easily (Bruemmer et al., 1988). Since then research has found that new Zn minerals are formed in soil, the ZneAl-layered double hydroxide (LDH), which may then gain silicate and become even less phytoavailable (Scheidegger et al., 1997). Formation of the Zn-LDH is favored at higher soil pH. Zn in long-term Zncontaminated soils is largely transformed to the Zn-LDH (Voegelin et al., 2005). Ni, Zn and Co form LDH species in contaminated soils (Voegelin et al., 2002), but Cd does not, helping to explain why Cd can remain adsorbed, labile and phytoavailable for decades after addition to soils (e.g., Kukier et al., 2010). There is some evidence that fertilizerapplied Cd becomes somewhat less phytoavailable over time (Hamon et al., 1998), but to a far lesser extent than Zn. One of the most important errors in research to better understand longterm risks from soil trace elements was considering that freshly added soluble metal salts represented long-term risk. Besides the long-term reactions of adsorbed metals to become less phytoavailable (hysterisis), other errors were identified. When a soluble salt, say ZnSO4, is added to soil, the Zn reacts with the soil surfaces and displaces protons causing lower pH and higher Zn phytoavailability (White et al., 1979; Speir et al., 1999). Large additions of Zn caused reduction in soil pH of several units which greatly increased the phytoavailability/phytotoxicity of the added Zn. Even if the soil is kept at the same pH by additions of CaCO3, until the anion is leached from the soil the presence of the accompanying anion causes soil solution to contain higher levels of Ca, Mg, etc. These compete with Zn for sorption sites and keep the activity of free Zn2þ in the soil solution higher. Thus until amended soils are leached to remove the accompanying anion, phytoavailability is higher whatever the pH. And until time has transpired to allow the Zn to move into new solid phases and sorption equilibria to be reached, Zn is much more phytoavailable than it becomes over time. Some of the best evidence to illustrate this point comes from study of Zn salt additions to soils collected near and far from galvanized steel towers for electricity lines. Near the towers, soils can reach 10,000 mg Zn kg
1. In strongly acidic soils this high Zn inhibits nitrification and causes phytotoxicity (which caused the selection/evolution of Zn tolerant grasses) (Al-Hiyaly et al., 1988). But when soil pH is near 6.5 or above for the very high Zn soils, no adverse effects are seen, not on plants or on soil microbes (Smolders et al. 2004, 2009). If equivalent Zn salt is added to the “control” soil collected far from the tower, and pH kept equivalent, severe toxicity to soil microbes was observed. Thus the extreme concern expressed by some about accumulation of Zn in soils being a threat to soil microbes is now seen as an artifact of the use of soluble Zn salts in study of microbial toxicity. These conclusions apply to all studies in which soluble metal salts are added to estimate effects of metals in soils.

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8. Other Elements in Fertilizers and Soil Amendments of Possible Concern Other elements occur in particular soil amendments which have raised concern about potential phytotoxicity or food-chain risk. These are considered below. In phosphates, F, As and U have raised concerns and risk assessment research has clarified the basis for risk.

8.1. Fluoride Phosphate fertilizers commonly contain considerable amounts of F (1e3% in superphosphate), and most of the F must be removed from P sources before such materials can be used as defluorinated phosphate feed additives. Excessive F absorption from contaminated feedstuffs or feed supplements by livestock interferes with bone and teeth health (fluorosis) (Suttie, 1980). The NRC (1980) suggests that the maximum allowable level of F in ruminant diets is about 40 mg kg
1 for the most sensitive grazing animal classes (bovines). Fortunately, F is strongly bound or precipitated in soils, is sufficiently poorly translocated to plant shoots, and is so phytotoxic that shoots of plants injured by soil F absorption do not exceed 40 mg kg
1 (Stevens et al., 2000). Most F poisoning of livestock occurred where crops were exposed to and accumulated high levels of F from air pollution of aluminum smelters, etc. (Weinstein, 1977; Weinstein and Davison, 2003). As might be expected, the bioavailability of different forms of F in diets vary considerably, with NaF more bioavailable than CaF2 than phosphate supplements (Shupe et al., 1962; Ammerman et al., 1964). Aluminum smelters or other industrial emissions commonly caused plant foliage to contain dangerous levels of F, but when orchardgrass was grown on the contaminated soil in a clean air greenhouse, the plants contained only basal levels of F (Braen and Weinstein, 1985). Besides aluminum smelters, phosphate fertilizer manufacturers, brick works, glass manufacturers, and electronics manufacturers commonly emit HF or SiF4 gases to the local atmosphere which can poison plants which can impact livestock and wildlife. Some steel smelting emits F, and downwind vegetation was highly F-enriched at a site in Utah. The forage F was fed to cattle in comparison with NaF and CaF2, and the forage F was found to be as toxic and bioavailable as NaF (Shupe et al., 1962). Concern about F uptake by forages grown on biosolids amended soils led to testing by Davis (1980) when a UK biosolids with 33500 mg F kg
1 was identified. Most biosolids contained only as high F as fertilized soils (about 300 mg kg
1) (Rea, 1979). Perennial ryegrass was grown on control and biosolids amended soil immediately after the biosolids were

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incorporated into the soil. Fluoride accumulation by three clippings of perennial ryegrass declined rapidly to near background levels after incorporation of the biosolids into the test soil. However, the potential for direct ingestion of such high F in biosolids amended surface soil, or biosolids adhering on plants, raises concerns. Chaney and Lloyd (1979) showed that fluid biosolids spread out on, dried on, and adhered to leaves unless immediately washed off. Adhering F-rich biosolids or manures could carry excessive F into diets of grazing livestock even when plants cannot accumulate excessive F from amended soils. Based on the evidence of livestock ingestion of biosolids adhering to forages, Kienholz et al. (1979) fed biosolids at 4 and 12% of diet to cattle to examine the bioavailability of biosolids trace elements and xenobiotics to livestock. Although the increase in bone F was not extreme and no adverse effects of the F were observed, animals fed biosolids did have significantly higher bone F than the control animals, and in that case, the F concentration in the biosolids was more typical at about 300 mg kg
1. Thus F-rich biosolids (enriched due to discharge by specific industries) should be injected or incorporated into soils and crop established before grazing livestock are exposed to a field. Highly F enriched biosolids should not be used in agriculture. To date, F in biosolids has not been regulated in the U.S. despite this evidence being provided to US-EPA. Risk assessment for F in phosphate fertilized soils has been conducted by several groups (Cronin et al., 2000; McLaughlin et al., 2001; Stewart et al., 1974; Loganathan et al., 2001, 2003, 2008; Grace et al., 2003, 2005). The initial risk assessments concluded that fertilizer-applied F could theoretically cause adverse effects in grazing cattle and sheep, and appreciable field research was conducted to support improved risk assessment. Eventually, soil feeding tests were conducted to assess soil-F bioavailability. Fluoride applied in phosphate fertilizers quickly reacts with soils to reach low phytoavailability (Loganathan et al., 2001; McLaughlin et al., 2001). Forages grown in fields with high cumulative loadings of phosphate borne F show a small increase in shoot F, but do not reach phytotoxic or zootoxic F levels in forages. One important aspect of the potential risk from fluoride ingestion by grazing livestock is the adherence of particles of the applied phosphate fertilizer on the surfaces of leaves (Stewart et al., 1974). Pasture herbage collected the day fertilizer was broadcast on the pasture contained up to 800 mg F kg
1, while unfertilized herbage contained < 2 mg F kg
1. By about 30 days post fertilization, F levels in the herbage dropped to background levels. Thus potential direct ingestion of the P-fertilizer adhering to leaves may require a livestock withholding period for P-fertilized pastures. In some field tests, when livestock grazed pastures with varied rates of phosphate fertilizer, the

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bone F of sheep increased significantly (but mildly) with increasing phosphate rate (Mason et al., 1989; Stewart et al., 1974). Now these experiments can be interpreted as direct ingestion of the phosphate fertilizer by the grazing animals which were not removed from the pastures during the experiment. But ingestion of soil by grazing livestock is well known, and the bioavailability of F in ingested surface soil of P-fertilized pastures (rich in F) was initially assumed to be high. Additional research evaluated the bioavailability of F in soils ingested by livestock, both in the field and in controlled feeding studies. Results have varied among studies from quite low to moderately high, based on different approaches to assessing “bioavailability” (e.g., Milhaud et al., 1989 vs. Grace et al., 2003). In the end, accumulation of F in bone is the key measure of risk from ingested F, not the net absorption of F from diets as reflected in urine F or balance studies. As noted above, CaF2 is substantially less bioavailable than NaF, so it should be no surprise that soil-F with its low solubility due to reaction of F with the soil minerals would have low F-bioavailability. Study of serum F in soil fed livestock indicated that until serum exceeded about 0.25 mg F L
1, no increase in bone F resulted. Urine F was increased more than bone F by soil feeding compared to NaF ingestion (Grace et al., 2003, 2005). Thus, soils with common levels of accumulated phosphate fertilizerF cause little increase in bone F in cattle or sheep. But crops enriched in soluble F by air pollution cause remarkable F accumulation in the animals and F disease. Even with soil containing 1000 mg F kg
1, no risk was evident from feeding research. Average soil ingestion is only about 2.5% of diet, while maximum short term soil ingestion reaches 15% of diet or higher (during winter with short pasture or over-stocked pastures). So the timing of exposure is important, especially in New Zealand research where winter pastures with inadequate forage causes increased soil ingestion (Grace et al., 2007). During periods when soil-F ingestion is low, some previously absorbed F is excreted by grazing livestock, so the long-term soil-F ingestion pattern shows lower risk than estimated from short term feeding studies at winter soil intake rates (Grace et al., 2007). Continuous exposure to high F soils is required to show clear evidence of higher F absorption, while livestock are rotated among pastures to maintain adequate forage supply and graze in seasons other than winter. And even with soils with relatively high accumulated F from high end phosphate fertilizer applications, little F was absorbed and deposited in bone during soil feeding tests. Because of the reactions of F in soils, plants and livestock, foods hardly cause any increase in exposure to F from soils with significant F enrichment except by soil ingestion. At many industrial/mining or F-mineralized sites, livestock and wildlife may be at risk from soil F (e.g., Geeson et al., 1998;

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Andrews and Cooke, 1989). Wildlife have suffered adverse effects on teeth and bones when exposed to high F soils even when plants accumulate little F internally (Schroder et al., 1999, 2003). Splash of high F soils onto plants or trampling of forages into soils by grazing livestock increase soil ingestionexposure of livestock and wildlife. An illustrative case of F poisoning of humans was identified in China where high F coal and charcoal were used to dry crops or heat homes in open fires which transferred ash particles very rich in F to foods and humans (Finkelman et al., 1999) Also, a few plant species strongly bioaccumulate F from soil; one of these is tea (Camilla sinensis). Several risk assessments have been conducted for F exposure from tea consumption (e.g., Lung et al., 2008). Phosphate fertilizers for tea probably should contain low levels of F to minimize F accumulation in tea.

8.2. Radionuclides Radionuclides (226Ra, 238U, 228Po and related isotopes) are commonly present in most phosphate rock ore (especially from marine sources compared to volcanic sources) (Menzel, 1968; Mortvedt, 1994). The radionuclides are distributed between products and wastes during processing with more 228Ra going to the phosphogypsum than the fertilizer. Electric furnace phosphate can be nearly devoid of radionuclides, while simple superphosphate contains most of the radionuclides in the rock. Phosphate fertilizers may increase soil U levels depending on their source and processing. Several kinds of risks from the radionuclides in fertilizers have been evaluated. Because 40K is a part of all natural K on Earth, humans receive radiation dose from the K in their bodies and dwellings. Risk assessments have been conducted for workers in the K fertilizer industry. The main conclusion of the 40K assessments is that one should not live in a house within a K fertilizer factory. Because full crop yield requires appropriate K fertilizer, and the K is a natural radiation source, all food has some 40 K. This is part of the background natural radiation dose of all on Earth. Fertilizers have little or no effect on 40K dose to humans; it is considered unavoidable and part of the normal and acceptable background radiation dose. Essentially all phosphates contain some U because phosphate ore bodies collect U from leaching groundwater (Menzel, 1968). Thus, phosphate fertilizers apply U to soils, and the accumulation of U in long-term soil fertility plots has been demonstrated (Rothbaum et al., 1979). Because of adsorption on the soil solid phases, U is effectively retained in the tillage depth. Because P fertilization increases U sorption by soils, adding phosphate fertilizer often decreases U concentrations in plants.

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Phosphogypsum storage piles, phosphate ore bodies and wastes, and P-fertilizer storage facilities generate radon gas (222Rn) which decomposes to 210Pb and comprises an avoidable radiation source. Most focus on 222 Rn risk is movement of geological 222Rn into dwellings, decomposition of the 222Rn into 210Pb, and exposure of humans through inhalation. Some 222 Rn dissolved in well water can provide human exposure through offgassing within homes, and drinking water (Milvy and Cothern, 1989). Surface water supplies contain much lower levels of 222Rn because it is degassed before or during water treatment. In areas with naturally high groundwater Ra, much of the Ra can be removed during water treatment to produce drinking water. Unfortunately, removal of the water Ra into the drinking water treatment residue makes this residue unacceptable for beneficial use in reducing phosphate in runoff of manured fields (Ippolito et al, 2011). It is estimated that 222Rn contributes about half of the background radiation to which humans are naturally exposed, and a local source of indoor air 222Rn (geological or groundwater) can substantially increase that risk (Cothern, 1999). Atmospheric 222Rn also decays to 210Pb which continuously falls onto Earth, including a minor contamination of all plant surfaces. The 210Pb from atmospheric sources is not considered a significant part of the natural background radiation exposures on Earth. But the 222Rn confined to a dwelling allows higher exposures which should be minimized. There is little evidence that fertilizer use contributes to Rn or Ra in groundwater, while phosphogypsum storage pile leachates can contribute 226Ra, but this is limited by the solubility of RaSO4 (Burnett and Elzerman, 2001), while use of high rates of phosphogypsum on cropland did not contribute to risk to leachate (Alcordo et al., 1999). Regulations on phosphogypsum are based on preventing excessive exposure to radon. U is very weakly absorbed by plants, even those growing in uranium mine wastes. Radium in fertilizer products is somewhat mobile in crops (Mortvedt, 1994). But little Ra moves to phloem fed storage tissues (seeds, fruits, storage roots and tubers) (Mays and Mortvedt, 1986). Leafy crops can accumulate some Ra from uranium or phosphate mine waste contaminated soils (Million et al., 1994; Tracy et al., 1983), but there is no evidence that normal fertilizer use increases Ra in foods (Sheppard et al., 1989). One interesting comparison in the literature is the evaluation of radionuclides in lettuce grown with nutrient solutions, or in manure fertilized and conventionally fertilized crops. Although much more U was added with the conventional fertilizer, fertilized lettuce radionuclides were no higher than the manured crop, and those were only slightly higher than the nutrient solution grown crop (Lauria et al., 2009). Lack of uptake and translocation of the fertilizer-derived radionuclides is evident in these results as well as those cited above.

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8.3. Arsenic A highly As and Pb contaminated Fe fertilizer was noted above (Williams et al., 2006), but the only common fertilizer with As concentrations higher than background soils are some phosphates. As levels in superphosphate range widely from 3e21 mg kg
1 (Table 1) (Chen et al., 2007) or up to 300 mg kg
1 in other fertilizers, while background US soils vary from <1.0e18 mg kg
1 (2e12 mg kg
1 for the 5the95th percentiles; Smith et al., 2005). Some livestock manure contains substantial levels of As from feed additives to promote growth or limit disease such that poultry litter and swine manure may commonly contain 25 mg As kg
1 DW (Nicholson et al., 1999). High soil As may also occur where Pb-arsenate was applied to orchards as a pesticide, or where arsenical herbicides or defoliants were repeatedly applied (Staed et al., 2009). The chemistry of As in soils is complicated by unequal adsorption of arsenate and arsenite, and the normal reduction of sorbed arsenate to arsenite in flooded soils. Weaker adsorption of arsenite than arsenate causes higher dissolved As in flooded soils, and hence higher potential for As uptake by rice than crops grown in aerobic soils. Rice is the crop species most sensitive to excessive soil As, which inhibits fertilization of flowers causing empty seeds (called straighthead disease) (Rahman et al., 2008). Interestingly, arsenate is normally absorbed by the phosphate transporter in roots (Zhao et al., 2009), but arsenite uptake is by a silicate transporter (Ma et al., 2008). Because As is so weakly accumulated by plants (Zhao et al., 2009, 2010), livestock and humans are principally at risk from As in ingested soil where the soil was contaminated by industrial pollution or geogenic enrichment. However, serious adverse human health effects of As in drinking water occurred in Taiwan, Bangladesh and West Bengal where drinking, cooking and irrigation water was highly enriched in As from deep geological materials rich in As (Kim et al., 2007). If flooded rice is irrigated with water containing high arsenate concentration, shoot and grain As are considerably increased (Abedin et al., 2002; Hossain et al., 2008). But if rice is grown on As enriched soils with low As irrigation water, rice is not so increased in As. And rice grown aerobically rather than by flooded culture accumulates significantly lower levels of As in leaves and grain (Arao et al., 2009; Li et al., 2009). Silicate amendment may also reduce As uptake by competition with arsenite (Li et al., 2009). As accumulation in rice grain has become a significant issue because risk estimations by several Agencies have led to suggestions that even background levels of As in uncontaminated soil comprise risk to children exposed to the soil. Some suggest that inorganic As in normal rice grown on uncontaminated soils is so high that efforts should be made to breed rice

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with lower As accumulation (Meharg and Raab, 2010; Zhao et al., 2010) or to advise against rice consumption. It does not seem likely that the most important human food crop on Earth (rice) which accumulates levels of only 0.2e0.4 mg As kg
1 in grain when grown on uncontaminated soils, should cause adverse health effects. Accumulation of As in the leaves and stem of rice is much higher than in the grain, so that rice stover might need to be avoided as livestock feed for long-lived breeding animals, but normal meat animals will not accumulate unusual levels of tissue inorganic As from such feedstuffs so the food-chain is protected from substantial increase in feed As. One interesting example of a problem feedstuffs and soil amendment rich in As is seaweed (Feldmann et al., 2000; Hansen et al., 2003). Where seaweed is available to farmers at low cost, it is often used as a feed or organic fertilizer. But because of the high As levels present in many seaweed species (Laminaria hyperborea and L. digitata averaged 74 mg As kg
1 DW) (Hansen et al., 2003), seaweed and manure from seaweed fed livestock should not be applied to soils. Organic forms of As in seaweed are rapidly converted to inorganic As in aerobic soils (Castlehouse et al., 2003), and may be bioavailable if the soil is ingested. Some fish and shellfish contain high levels of As mostly as arsenobetaine. This organo-As chemical is believed to comprise no risk to consumers because it is absorbed and excreted unchanged and does not cause adverse effects in the body (Vahter et al., 1983). But some of the As in these foods is inorganic (Buchet et al., 1994), so ingestion of fish and shellfish may provide more inorganic As than the vegetable foods (except rice) in the normal diet. Interestingly, some cultures ingest significant amounts of seaweed or cooked seaweed (e.g., hijiki which provides 23 mg inorganic As per meal; Nakamura et al., 2008). Because of As accumulation in seaweed, fish and shellfish, composted residues of these foods may need to be avoided for land application to limit adding As to soils. Another possible aspect of As in agriculture is the contamination of soil by As in CCA-treated lumber used to construct raised beds for gardening (Rahman et al., 2004). Although the As, Cu and Cr do not leach far from the wood, soil within the 0e2 cm zone from the wood surface with original 3e10 mg As kg
1 was raised to 50 mg As kg
1 after a few years. Contaminated soil from the 0e2 cm zone was used to grow test garden crops and As was slightly but significantly increased in the edible portions of carrot, spinach, etc. Carrot peel had higher As than the peeled carrot. Spinach reached 1.48 mg As kg
1 DW, while control spinach contained only about 0.070 mg kg
1 DW. Although the increase in crop As grown in soil very near the treated wood is appreciable, the principal As exposure from CCA treated wood is soil/ dust ingestion by children who touch the wood or the soil adjacent to the wood.

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Unfortunately, the risk assessments conducted by US-EPA have used faulty data in the dose response estimation for cancer risk (see Chappell et al., 1997). In Taiwan, residents of a village might have consumed low As water from a surface well, or very high As water from a deep artesian well. But when the cancer incidence epidemiological evaluation was conducted, the data for well water As for different persons was not available. So they used the village median well As concentration to classify all persons in a village, miscategorizing persons with cancer in the village who could have had either high or low water As. Other flaws in the epidemiology have been reported (Chappell et al., 1997; Lamm et al., 2006). In the US, several studies have examined persons who consumed water with moderately high As for decades with no evidence of the bladder cancers caused by excessive As from highly contaminated water in Taiwan (Meliker et al., 2010; Lamm et al., 2004). This is strong evidence that a threshold exists for water As concentration below which no cancer risk exists. Intense policy debate about the risk from low level As exposure continues. During 2010, US-EPA proposed to increase the cancer slope factor for As by 17-fold (US-EPA, 2010). This would reduce the allowed As in drinking water to levels which cannot practically be reached. But worse, the cancer based Soil Screening Level which has been 0.43 mg kg
1 (already below background soil As levels) would be lowered to 0.025 mg kg
1, far below the natural level in all soils (normal 5the95th percentile range, 2e12 mg kg
1) (Smith et al., 2005). Such low possible limits are not a problem for US drinking water regulations because they set the “Maximum Contaminant Level Goal” at 0 mg As L
1, and then set a regulatory limit based on practical matters of achieving and measuring the water As. Public concern about historic soil As contamination risk has been raised by increasing knowledge about soil As from gold mine wastes, Cu and Sn mine wastes, pesticides, defoliants, CCA-treated wood products and other As sources. On the other hand, if soils are rich in amorphous Fe, the bioavailability of soil-As is considerably reduced by formation of FeAsO4 (scorodite) or AsO24 adsorbed on Fe oxides (Scheckel et al., 2009; Meunier et al., 2010; Beak et al., 2006). The implications for As in fertilizers and soil amendments of such high cancer slope factors for As are complex. Final decisions may not be reached for years because of lawsuits filed to challenge the change in cancer slope factor. All phosphates contain some As, and if the soil As limit is below background, As in any fertilizer product could be challenged. Such a conclusion seems irrational; increasing soil As by a few percent over decades of fertilizer application can hardly comprise the increase in risk being suggested by toxicologists. In two cases in the US, aerosol emissions of As from industry caused local soils where children lived to become strongly contaminated with

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As. In order to evaluate the potential risk to these children, their exposures were assessed, local soil As levels measured, and speciated As in children's urine was analyzed to assess actual individual inorganic As intakes (Hwang et al., 1997; Tsuji et al., 2005). Until soil exceeded 40e100 mg As kg
1, there was no evidence of increased inorganic As ingestion based on As species in children's urine. Whether this low As in urine should be attributed to lower soil ingestion by the children than assumed, or lower bioavailability of soil As, remains unknown. Several soil As bioavailability feeding tests have been reported using monkeys (Freeman et al., 1995; Roberts et al., 2007; Roberts et al., 2002) or pigs ( Juhasz et al., 2007; Rodriguez et al., 2003) which showed that soil As bioavailability was typically 15% of total, far lower than the 100% bioavailable assumed in current risk assessments. For some soils rich in amorphous Fe oxides, bioavailability has been as low as 2% of total As.

9. Monitoring and Control of Trace Elements in Mineral Fertilizers Based on the experiences and research summarized above, it is clear that fertilizer and soil amendment products with substantial trace element (Cd, As, Zn, Cu, Pb, Mo, Se) enrichment may not be acceptable for long-term use on agricultural soils. After the 1997 Seattle Times articles about contaminated fertilizers (see Wilson, 2001), many jurisdictions developed limits for metals in fertilizers. Several quite different approaches have been used, to varied effect in reducing Cd, Pb and As in fertilizers. Many felt that Cd limits should be based on the P (or P2O5) levels in the product because most of the Cd in P and mixed fertilizers comes from rock phosphate used to produce the fertilizer. But research has shown that high Cd products may cause clear increase in crop Cd over decades of use (Williams and David, 1973; McLaughlin et al., 1996; Grant and Sheppard, 2009). Questions were raised above about the levels of Cd which are “potentially tolerable” in human diets because of different views on which dietary Cd levels caused the first adverse effects on humans in a lifetime, the bioavailability of Cd in different foods, and the role of crop Zn inhibiting crop Cd absorption by humans. But confusion about the basis for limits has prevented the needed controls from being established in many jurisdictions. Clearly, some limit on Cd in phosphates is needed. Table 2 shows the present limits for a 45% P2O5 product (triplesuperphosphate) in different control schemes. Because some jurisdictions limit annual or 45-year cumulative applications, one must assume a P2O5 application rate to estimate the limit for Cd in the product, so in the present comparison,

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100 kg P2O5 ha
1 yr
1 was assumed (say for potatoes, much higher rate than for wheat). Based on the review above, it is appears that the EU proposal of 20 mg Cd kg
1 P2O5 may be unnecessarily low, and the AAPFCO proposed 450 mg Cd kg
1 P2O5 may be too high to limit crop Cd in the long term based on the Australian experience. The cocontamination with Zn reduces risk from Cd in some products, but this is not considered in present proposals or rules. Some phosphate ore reserves will require Cd removal before products should be marketed. But Cd-free products are not required to protect public health over centuries of risk consideration, perhaps longer than known P reserves are predicted to last (note above discussion of lack of increase in crop Cd level over time despite continuing use of P-fertilizers and other Cd sources for cropland).

10. Need for Regulatory Enforcement on Composition of Soil Amendments The contaminated ZnSO4 fertilizer case discussed above is a valuable example of why regulatory controls are needed. In the USA, States regulate fertilizer products; initially to assure adequate levels of the purchased nutrients, and later to assure lack of contamination. Representatives of the 50 States work thru the AAPFCO to develop the basis for uniform regulations which would be enforced at the State level. Uniform regulations support interstate commerce and are necessary in a large nation. In most US States, the Department of Agricultural has staff who actually sample products around their state which is the reason companies need to comply with the rules. Companies have to analyze potential products and submit data to obtain a license to sell fertilizer products, and they have to analyze their products often and report at least annually to the State governments where they market. Products must comply with the State rules or they may not be marketed. Other than Cd in phosphates, few commercial fertilizer products have ever exceeded regulated levels of trace elements. But some garden and lawn fertilizers have contained very high levels of Pb and As and some other contaminants. Fortunately when the product “Ironite” was initially marketed with 4000 mg kg
1 of As and Pb, States could apply their rules. Some states prohibited sale while other States required a package listed application rates to low that the product would not provide any benefit. When they were unable to obtain permits to market their product, the company found a different source of mining byproduct which did not have such high contamination with Pb and As or other unwanted materials. Contrast this with the international sale of the Cd-contaminated Zn fertilizer product described above. Some nations had essentially no

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impediment to open marketing of the product. And other nations with good rules did not have the ability to require pre-analysis and permit issuance before the product could be sold for use as a fertilizer or feed ingredient e Even in the EU and Australia. Control or regulatory programs are needed to limit this kind of adulterated fertilizer product from being marketed anywhere, but especially in lands were subsistence rice farming is common. Effective rule enforcement can prevent any of the risks discussed in this paper from happening in practice.

REFERENCES Abedin, M. J., Cotter-Howells, J., and Meharg, A. A. (2002). Arsenic uptake and accumulation in rice (Oryza sativa L.) irrigated with contaminated water. Plant Soil 240, 311e319. Adeli, A., Sistani, K. R., Tewolde, H., and Rowe, D. E. (2007). Broiler litter application effects on selected trace elements under conventional and no-till systems. Soil Sci. 172, 349e365. Åkesson, A., Lundh, T., Vahter, M., Bjellerup, P., Lidfeldt, J., Nerbrand, V., Samsioe, G., Str omberg, U., and Skerfving, S. (2005). Tubular and glomerular kidney effects in Swedish women with low environmental cadmium exposure. Environ. Health Perspect. 113, 1627e1631. Alcordo, I. S., Rechcigl, J. E., Roessler, C. E., and Littell, R. C. (1999). Radiological impact of phosphogypsum applied to soils under bahiagrass pasture. J. Environ. Qual. 28, 1555e1567. Alfven, T., Elinder, C. G., Carlsson, M. D., Grubb, A., Hellstrom, L., Persson, B., Pettersson, C., Spang, G., Schutz, A., and Jarup, L. (2000). Low-level cadmium exposure and osteoporosis. J. Bone Miner. Res. 15, 1579e1586. Al-Hiyaly, S. A. K., McNeilly, T., and Bradshaw, A. D. (1988). The effect of zinc contamination from electricity pylons: Evolution in a replicated situation. New Phytol. 110, 571e580. Ammerman, C. B., Arrington, R. L., Shirley, R. L., and Davis, G. K. (1964). Comparative effects of fluorine from soft phosphate, calcium fluoride and sodium fluoride on steers. J. Anim. Sci. 23, 409e413. Andersson, A., and Bingefors, S. (1985). Trends and annual variations in Cd concentrations in grain of winter wheat. Acta Agric. Scand. 35, 339e344. Andersson, A., and Hahlin, M. (1981). Cadmium effects from phosphorus fertilization in field experiments. Swed. J. Agric. Res. 11, 3e10. Andrews, S. M., and Cooke, J. A. (1989). Distribution of trace element pollutants in a contaminated ecosystem established on metalliferous fluorspar tailings. 3: Fluoride. Environ. Pollut. 60, 165e179. Aoshima, K. (1987). Epidemiology of renal tubular dysfunction in the inhabitants of a cadmium-polluted area in the Jinzu River basin in Toyama Prefecture. Tohoku J. Exp. Med. 152, 151e72. Aoshima, K., Kawanishi, Y., Fan, J. J., Cai, Y. Q., Katoh, T., Teranishi, H., and Kasuya, M. (1995). Cross-sectional assessment of renal function in the inhabitants of a cadmiumpolluted area. Ann. Clin. Lab. Sci. 25, 493e503. Arao, T., Kawasaki, A., Baba, K., Mori, S., and Matsumoto, S. (2009). Effect of water management on cadmium and arsenic accumulation and dimethylarsinic acid concentrations in Japanese rice. Environ. Sci. Technol. 43, 9361e9367. Arthur, M. A., Rubin, G., Woodbury, P. B., Schneider, R. E., and Weinstein, L. H. (1992). Uptake and accumulation of selenium by terrestrial plants growing on a coal fly ash landfill: Part 2. Forage and root crops. Environ. Toxicol. Chem. 11, 1289e1299.

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Bannon, D. I., Abounader, R., Lees, P. S., and Bressler, J. P. (2003). Effect of DMT1 knockdown on iron, cadmium, and lead uptake in Caco-2 cells. Am. J. Physiol. Cell. Physiol. 284, C44eC50. Bañuelos, G. S. (2006). Phyto-products may be essential for sustainability and implementation of phytoremediation. Environ. Pollut. 144, 19e23. Basta, N. T., Ryan, J. A., and Chaney, R. L. (2005). Trace element chemistry in residualtreated soil: Key concepts and metal bioavailability. J. Environ. Qual. 34, 49e63. Beak, D. G., Basta, N. T., Scheckel, K. G., and Traina, S. J. (2006). Bioaccessibility of arsenic (V) bound to ferrihydrite using a simulated gastrointestinal system. Environ. Sci. Technol. 40, 1364e1370. Bell, P. F., Mulchi, C. L., and Chaney, R. L. (1988). Residual effects of land applied municipal sludge on tobacco. III. Agronomic, chemical, and physical properties vs. multiple sludge sources. Tobacco Sci. 32, 71e76, (Tobacco Int. 190, 47e52.). Bell, P. F., Parker, D. R., and Page, A. L. (1992). Contrasting selenate-sulfate interactions in selenium-accumulating and nonaccumulating plant species. Soil Sci. Soc. Am. J. 56, 1818e1824. Bengtsson, H., Öborn, I., Jonsson, S., Nilsson, I., and Andersson, A. (2003). Field balances of some mineral nutrients and trace elements in organic and conventional dairy farming e A case study at Öjebyn, Sweden. Eur. J. Agron. 20, 101e116. Bingham, F. T., Sposito, G., and Strong, J. E. (1984). The effect of chloride on the availability of cadmium. J. Environ. Qual. 13, 71e74. Bingham, F. T., Sposito, G., and Strong, J. E. (1986). The effect of sulfate on the availability of cadmium. Soil Sci. 141, 172e177. Boawn, L. C. (1976). Sequel to “residual availability of fertilizer zinc”. Soil Sci. Soc. Am. J. 40, 467e468. Braen, S. N., and Weinstein, L. H. (1985). Uptake of fluoride and aluminum by plants grown in contaminated soils. Water Air Soil Pollut. 24, 215e223. Bramley, R. G. V., and Barrow, N. J. (1994). Differences in the cadmium content of some common Western Australian pasture plants grown in a soil amended with cadmium e Describing the effects of level of cadmium supply. Fert. Res. 39, 113e122. Bray, B. J., Dowdy, R. H., Goodrich, R. D., and Pamp, D. E. (1985). Trace metal accumulations in tissues of goats fed silage produced on sewage sludge-amended soil. J. Environ. Qual. 14, 114e118. Broadley, M. R., Alcock, J., Alford, J., Cartwright, P., Foot, I., Fairweather-Tait, S. J., Hart, D. J., Hurst, R., Knott, P., McGrath, S. P., Meacham, M. C., Norman, K., Mowat, H., Scott, P., Stroud, J. L., Tovey, M., Tucker, M., White, P. J., Young, S. D., and Zhao, F.-J. (2010). Selenium biofortification of high-yielding winter wheat (Triticum aestivum L.) by liquid or granular Se fertilisation. Plant Soil 332, 5e18. Bruemmer, G. W., Gerth, J., and Tiller, K. G. (1988). Reaction kinetics of the adsorption and desorption of nickel, zinc, and cadmium by goethite. I. Adsorption and diffusion of metals. J. Soil Sci. 39, 37e52. Buchet, J. P., Pauwels, J., and Lauwerys, R. (1994). Assessment of exposure to inorganic arsenic following ingestion of marine organisms by volunteers. Environ. Res. 66, 44e51. Burau, R. G. (1983). National and local dietary impact of cadmium in south coastal California soils. Ecotoxicol. Environ. Saf. 7, 53e57. Burnett, W. C., and Elzerman, A. W. (2001). Nuclide migration and the environmental radiochemistry of Florida phosphogypsum. J. Environ. Radioact. 54, 27e51. Cai, S., Yue, L., Hu, Z., Zhong, X., Ye, Z., Xu, H., Liu, Y., Ji, R., Zhang, W., and Zhang, F. (1990). Cadmium exposure and health effects among residents in an irrigation area with ore dressing wastewater. Sci. Total Environ. 90, 67e73.

Food Safety Issues for Mineral and Organic Fertilizers

101

Castlehouse, H., Smith, C., Raab, A., Deacon, C., Meharg, A. A., and Feldmann, J. (2003). Biotransformation and accumulation of arsenic in soil amended with seaweed. Environ. Sci. Technol. 37, 951e957. CDFA (California Department of Food and Agriculture). (1998). Development of Risk-based Concentrations for Arsenic, Cadmium, and Lead in Inorganic Commercial Fertilizers. Sacramento, CA: Foster-Wheeler Environmental Corporation. Chaney, R. L. (1983). Potential effects of waste constituents on the food chain. In J. F. Parr, P. B. Marsh, and J. M. Kla (Eds.), Land Treatment of Hazardous Wastes (pp. 152e240). Park Ridge, NJ: Noyes Data Corp. Chaney, R. L. (1993). Zinc phytotoxicity. In A. D. Robson (Ed.), Zinc in Soils and Plants (pp. 135e150). Dordrecht: Kluwer Academic Publ. Chaney, R. L. (2010). Cadmium and zinc. Chapter 17. In P. Hooda (Ed.), “Trace Elements in Soils.” (pp. 409e439). Oxford, UK: Blackwell Publ. Chaney, R. L., Beyer, W. N., Gifford, C. H., and Sileo, L. (1988). Effects of zinc smelter emissions on farms and gardens at Palmerton, PA. Trace Subst. Environ. Health 22, 263e280. Chaney, R. L., Broadhurst, C. L., and Centofanti, T. (2010a). Phytoremediation of soil trace elements. Chapter 17. In P. Hooda (Ed.), “Trace Elements in Soils.” (pp. 311e352). Oxford, UK: Blackwell Publ. Chaney, R. L., Codling, E. E., Scheckel, K. G., and Zia, M. H. (2010b). Pb in Carrots Grown on Pb-Rich Soils is Mostly Within the Xylem. Abstract No. 60451, American Society of Agronomy Annual Meeting (Long Beach, CA, Oct. 31 to Nov. 4, 2010). Madison, WI: Amer. Soc. Agron. Chaney, R. L., Filcheva, E., Green, C. E., and Brown, S. L. (2006). Zn deficiency promotes Cd accumulation by lettuce from biosolids amended soils with high Cd:Zn ratio. J. Residual Sci. Technol. 3, 68e75. Chaney, R. L., Green, C. E., Ajwa, H., and Smith, R. (2009). Zinc Fertilization Plus Liming to Reduce Cadmium Uptake by Romaine Lettuce on Cd-Mineralized Lockwood Soil. http:// repositories.cdlib.org/ipnc/xvi/1252 Proc. Int. Plant Nutrition Colloquium XVI (Aug. 25e28, 2009, Sacramento, CA): Paper 1252. Chaney, R. L., and Lloyd, C. A. (1979). Adherence of spray-applied liquid sewage sludge to tall fescue. J. Environ. Qual. 8, 407e411. Chaney, R. L., Norton, L. D., Dayton, E. L., Torbert, H. A., Schomberg, H. H., Codling, E. E., Dick, W. A., Taylor, T., and Smith, M. C. (2010c). Beneficial Use of FGD-Gypsum in Agriculture: How Uncertainty in Risk Assessment and Modeling Affects Policy. Abstract No. 60453, American Society of Agronomy Annual Meeting (Long Beach, CA, Oct. 31 to Nov. 4, 2010). Madison, WI: Amer. Soc. Agron. Chaney, R. L., and Ryan, J. A. (1993). Heavy metals and toxic organic pollutants in MSWcomposts: Research results on phytoavailability, bioavailability, etc. In H. A. J. Hoitink, and H. M. Keener (Eds.), Science and Engineering of Composting: Design, Environmental, Microbiological and Utilization Aspects (pp. 451e506) Columbus, OH: Ohio State Univ. Chaney, R. L., and Ryan, J. A. (1994). Risk Based Standards for Arsenic, Lead and Cadmium in Urban Soils (ISBN 3-926959-63-0). DECHEMA, Frankfurt: Germany. Chaney, R. L., Ryan, J. A., Kukier, U., Brown, S. L., Siebielec, G., Malik, M., and Angle, J. S. (2001). Heavy metal aspects of compost use. In P. J. Stofella, and B. A. Kahn (Eds.), “Compost Utilization in Horticultural Cropping Systems.” (pp. 323e359). Boca Raton, FL: CRC Press. Chaney, R. L., Stoewsand, G. S., Furr, A. K., Bache, C. A., and Lisk, D. J. (1978). Elemental content of tissues of Guinea pigs fed Swiss chard grown on municipal sewage sludge-amended soil. J. Agr. Food Chem. 26, 944e997. Chappell, W. R., Beck, B. D., Brown, K., Chaney, R. L., Cothern, R., Irgolic, K., North, W., Thornton, I., and Tsongas, T. (1997). Inorganic arsenic: A need and an opportunity to improve risk assessment. Environ. Health Perspect. 105, 1060e1067.

102

Rufus L. Chaney

Chaudri, A. M., Zhao, F.-J., McGrath, S. P., and Crosland, A. R. (1995). The cadmium content of British wheat grain. J. Environ. Qual. 24, 850e855. Chen, W., Chang, A. C., and Wu, L. (2007). Assessing long-term environmental risks of trace elements in phosphate fertilizers. Ecotox. Environ. Safe. 67, 48e58. Chen, W., Li, L., Chang, A. C., Wu, L., Chaney, R. L., Smith, R., and Ajwa, H. (2009). Characterizing the solid-solution partitioning coefficient and plant uptake factor of As, Cd, and Pb in California croplands. Agric. Ecosyst. Environ. 129, 212e220. Chien, S. H., and Menon, R. G. (1994). Uptake of phosphorus and cadmium by maize from an acid soil treated with high rates of phosphate fertilizers varying in cadmium content. In D. C. Adriano, Z.-S. Chen, and S.-S Yang (Eds.), Biogeochemistry of Trace Elements (pp. 205e214). London, UK: Science and Technol. Lett. CODEX (Codex Alimentarius Commission). (2006). Report of the 29th Session of the Codex Alimentarius Commission. ALINORM 06/29/41. Rome: Codex Alimentarius Commission. http://www.codexalimentarius.net/web/archives.jsp?year¼06. Codling, E. E., Chaney, R. L., and Mulchi, C. L. (2008). Effects of broiler litter management practices on phosphorus, copper, zinc, manganese and arsenic concentrations in Maryland Coastal Plain soils. Commun. Soil Sci. Plant Anal. 39, 1193e1205. Codling, E. E., Green, C. E., and Chaney, R. L. (2012) Accumulation of lead and arsenic by carrots grown on four lead-arsenate contaminated orchard soils. Submitted to J. Plant Nutr. Cohen, C. K., Garvin, D. F., and Kochian, L. V. (2004). Kinetic properties of a micronutrient transporter from Pisum sativum indicate a primary function in Fe uptake from the soil. Planta 218, 784e792. Copes, R., Clark, N. A., Rideout, R., Palaty, J., and Teschke, K. (2008). Uptake of cadmium from Pacific oysters (Crassostrea gigas) in British Columbia oyster growers. Environ. Res. 107, 160e169. Cothern, C. R. (1999). Indoor air radon. Environ. Geochem. Health 21, 83e90. Cronin, S. J., Manoharan, V., Hedley, M. J., and Loganathan, P. (2000). Fluoride: A review of its fate, bioavailability, and risks of fluorosis in grazed-pasture systems in New Zealand. N. Z. J. Agr. Res. 43, 295e321. CSTEE (Scientific Committee on Toxicity, Ecotoxicity and The Environment). (2002). Opinion of the CSTEE ON “Member State Assessments of the Risk to Health and the Environment from Cadmium in Fertilizers” Brussels. http://ec.europa.eu/enterprise/sectors/ chemicals/files/cadmium/sctee_en.pdf. Cupit, M., Larsson, O., de Meeûs, C., Eduljee, G. H., and Hutton, M. (2002). Assessment and management of risks arising from exposure to cadmium in fertilisers e II. Sci. Total Environ. 291, 189e206. Davies, B. E., Paveley, C. F., and Wixson, B. G. (1993). Use of limestone wastes from metal mining as agricultural lime: Potential heavy metal limitations. Soil Use Manag. 9, 47e52. Davis, J. G., Weeks, G., and Parker, M. B. (1995). Use of deep tillage and liming to reduce zinc toxicity in peanuts grown on flue dust contaminated land. Soil Technol. 8, 85e89. Davis, R. D. (1980). Uptake of fluoride by ryegrass grown in soil treated with sewage sludge. Environ. Pollut. B1, 277e284. de Burbure, C., Buchet, J. P., Bernard, A., Leroyer, A., Nisse, C., Haguenoer, J. M., Bergamaschi, E., and Mutti, A. (2003). Biomarkers of renal effects in children and adults with low environmental exposure to heavy metals. J. Toxicol. Environ. Health A66, 783e798. Decker, A. M., Chaney, R. L., Davidson, J. P., Rumsey, T. S., Mohanty, S. B., and Hammond, R. C. (1980). Animal performance on pastures topdressed with liquid sewage sludge and sludge compost. In Proc. Nat. Conf. Municipal and Industrial Sludge Utilization and Disposal (pp. 37e41). Silver Spring, MD: Information Transfer, Inc.

Food Safety Issues for Mineral and Organic Fertilizers

103

Degryse, F., Smolders, E., and Merckx, R. (2006a). Labile Cd complexes increase Cd availability to plants. Environ. Sci. Technol. 40, 830e836. Degryse, F., Smolders, E., and Parker, D. R. (2006b). An agar gel technique demonstrates diffusion limitations to cadmium uptake by higher plants. Environ. Chem. 3, 419e423. de Meeûs, C., Eduljee, G. H., and Hutton, H. (2002). Assessment and management of risks arising from exposure to cadmium in fertilisers. I. Sci. Total Environ. 291, 167e187. deVries, M. P. C., and Tiller, K. G. (1978). Sewage sludge as a soil amendment, with special reference to Cd, Cu, Mn, Ni, Pb, and Zn e Comparison of results from experiments conducted inside and outside a greenhouse. Environ. Pollut. 16, 213e240. EFSA (European Food Safety Agency). (2009). Scientific opinion: Cadmium in food; Scientific Opinion of the Panel on Contaminants in the Food Chain. EFSA J. 980, 1e139. http://www.efsa.europa.eu/en/efsajournal/doc/980.pdf. EFSA (European Food Safety Agency). (2010). Statement on tolerable weekly intake of cadmium. EFSA J. 9(2). http://www.efsa.europa.eu/en/efsajournal/doc/1975.pdf 1975. Ekholm, P., Reinivuo, H., Mattila, P., Pakkala, H., Koponen, J., Happonen, A., Hellstrom, J., and Ovaskainen, M. L. (2007). Changes in the mineral and trace element contents of cereals, fruits and vegetables in Finland. J. Food Comp. Anal. 20, 487e495. Elfving, D. C., Haschek, W. M., Stehn, R. A., Bache, C. A., and Lisk, D. J. (1978). Heavy metal residues in plants cultivated on and in small mammals indigenous to old orchard soils. Arch. Environ. Health 33, 95e99. Elinder, C.-G., Kjellström, T., Friberg, L., Lind, B., and Linmann, L. (1976). Cadmium in kidney cortex, liver, and pancreas from Swedish autopsies: Estimation of biological half time in kidney cortex, considering calorie intake and smoking habits. Arch. Environ. Health 31, 292e302. Ellis, K. J., Vartsky, D., Zanzi, I., and Cohn, S. H. (1979). Cadmium: In vivo measurement in smokers and nonsmokers. Science 205, 323e325. Ely, R. E., Dunn, K. M., and Huffman, C. F. (1948). Cobalt toxicity in calves resulting from high oral administration. J. Anim. Sci. 7, 239e246. EU (European Union). (1988). Draft Proposal Relating to Cadmium in Fertilizers Chemicals Unit of DG Enterprise. http://ec.europa.eu/enterprise/newsroom/cf/_getdocument.cfm?doc_ id¼2966. Ewers, U., Brockhaus, A., Dolgner, R., Freier, I., Jermann, E., Bernard, A., StillerWinkler, R., Hahn, R., and Manojlovic, N. (1985). Environmental exposure to cadmium and renal function of elderly women living in cadmium-polluted areas of the Federal Republic of Germany. Int. Arch. Occup. Environ. Health 55, 217e239. Feldmann, J., John, K., and Pengprecha, P. (2000). Arsenic metabolism in seaweed-eating sheep from Northern Scotland. Fres. J. Anal. Chem. 368, 116e121. Finkelman, R. B., Belkin, H. E., and Zheng, B.-S. (1999). Health impacts of domestic coal use in China. Proc. Natl. Acad. Sci. U.S.A. 96, 3427e3431. Flanagan, P. R., McLellan, J. S., Haist, J., Cherian, M. G., Chamberlain, M. J., and Valberg, L. S. (1978). Increased dietary cadmium absorption in mice and human subjects with iron deficiency. Gastroenterol. 74, 841e846. Fox, M. R. S., Jacobs, R. M., Jones, A. O. L., and Fry, B. E., Jr. (1979). Effects of nutritional factors on metabolism of dietary cadmium at levels similar to those of man. Environ. Health Perspect. 28, 107e114. Freeman, G. B., Schoof, R. A., Ruby, M. V., Davis, A. O., Dill, J. A., Liao, S. C., Lapin, C. A., and Bergstrom, P. D. (1995). Bioavailability of arsenic in soil and house dust impacted by smelter activities following oral administration in Cynomolgus monkeys. Fund. Appl. Toxicol. 28, 215e222. Friberg, L., Elinder, C. G., Kjellstrom, T., and Nordberg, G. F. (1985). Cadmium and Health: A Toxicological and Epidemiological AppraisalIn Exposure, Dose, and Metabolism Vol. 1. Boca Raton, FL: CRC Press.

104

Rufus L. Chaney

Friis, L., Petersson, L., and Edling, C. (1998). Reduced cadmium levels in human kidney cortex in Sweden. Environ. Health Perspect. 106, 175e178. Fukushima, M., Ishizaki, A., Sakamoto, M., and Kobayashi, E. (1973). Cadmium concentration in rice eaten by farmers in the Jinzu River basin (in Japanese). Japan J. Hyg. 28, 406e415. Gao, X., Brown, K. R., Racz, G. J., and Grant, C. A. (2010). Concentration of cadmium in durum wheat as affected by time, source and placement of nitrogen fertilization under reduced and conventional-tillage management. Plant Soil 337, 341e354. Garrett, R. G., Porter, A. R. D., Hunt, P. A., and Lalor, G. C. (2008). The presence of anomalous trace element levels in present day Jamaican soils and the geochemistry of Late-Miocene or Pliocene phosphorites. Appl. Geochem. 23, 822e834. Geeson, N. A., Abrahams, P. W., Murphy, M. P., and Thornton, I. (1998). Fluorine and metal enrichment of soils and pasture herbage in the old mining areas of Derbyshire, UK. Agric. Ecosyst. Environ. 68, 217e231. Gould, D. H. (1998). Polioencephalomalacia. J. Anim. Sci. 76, 309e314. Grace, N. D., Loganathan, P., Deighton, M. W., Molano, G., and Hedley, M. J. (2005). Ingestion of soil fluorine: Its impact on the fluorine metabolism of dairy cows. N. Z. J. Agr. Res. 48, 23e27. Grace, N. D., Loganathan, P., and Hedley, M. J. (2007). The effect of a temporal change in ingestion rates of fluorine (F) in soil on the concentration of F in serum and bone of young sheep. N. Z. Vet. J. 55, 77e80. Grace, N. D., Loganathan, P., Hedley, M. J., and Wallace, G. C. (2003). Ingestion of soil fluorine: Its impact on the fluorine metabolism and status of grazing young sheep. N. Z. J. Agr. Res. 46, 279e286. Grant, C. A., Bailey, L. D., Harapiak, J. T., and Flore, N. A. (2002). Effect of phosphate source, rate and cadmium content and use of Penicillium bilaii on phosphorus, zinc and cadmium concentration in durum wheat grain. J. Sci. Food Agric. 82, 301e308. Grant, C. A., Bailey, L. D., McLaughlin, M. J., and Singh, B. R. (1999). Management factors which influence cadmium concentrations in crops: A review. In McLaughlin, and B. R. Singh (Eds.), Cadmium in Soils and Plants (pp. 151e198). Dordrecht: Kluwer Academic Publ. Grant, C. A., Clarke, J. M., Duguid, S., and Chaney, R. L. (2008). Selection and breeding of plant cultivars to minimize cadmium accumulation. Sci. Total Environ. 390, 301e310. Grant, C. A., Dribnenki, J. C. P., and Bailey, L. D. (2000). Cadmium and zinc concentrations and ratios in seed and tissue of solin (cv LinolaÔ 947) and flax (cvs McGregor and Vimy) as affected by nitrogen and phosphorus fertiliser and provide (Penicillium bilaji ). J. Sci. Food Agric. 80, 1735e1743. Grant, C. A., Monreal, M. A., Irvine, R. B., Mohr, R. M., McLaren, D. L., and Khakbazan, M. (2010). Preceding crop and phosphorus fertilization affect cadmium and zinc concentration of flaxseed under conventional and reduced tillage. Plant Soil 333, 337e350. Grant, C. A., and Sheppard, S. C. (2008). Fertilizer impacts on cadmium availability in agricultural soils and crops. Human Ecol. Risk Assess. 14, 210e228. Gutenmann, W. H., Pakkala, I. S., Churey, D. J., Kelly, W. C., and Lisk, D. J. (1979). Arsenic, boron, molybdenum, and selenium in successive cuttings of forage crops grown on fly ash amended soil. J. Agr. Food Chem. 27, 1393e1395. Hamon, R. E., McLaughlin, M. J., Naidu, R., and Correll, R. (1998). Long-term changes in cadmium bioavailability in soil. Environ. Sci. Technol. 32, 3699e3703. Hamon, R., Wundke, J., McLaughlin, M., and Naidu, R. (1997). Availability of zinc and cadmium to different plant species. Aust. J. Soil Res. 35, 1267e1277.

Food Safety Issues for Mineral and Organic Fertilizers

105

Hansen, H. R., Raab, A., Francesconi, K. A., and Feldmann, J. (2003). Metabolism of arsenic by sheep chronically exposed to arsenosugars as a normal part of their diet. 1. Quantitative intake, uptake, and excretion. Environ. Sci. Technol. 37, 845e851. Harcz, P., De Temmerman, L., De Voghel, S., Waegeneers, N., Wilmart, O., Vromman, V., Schmit, J. F., Moons, E., Van Peteghem, C., De Saeger, S., Schneider, Y. J., Larondelle, Y., and Pussemier, L. (2007). Contaminants in organically and conventionally produced winter wheat (Triticum aestivum) in Belgium. Food Add. Contam. 24, 713e720. Hart, J. J., Welch, R. M., Norvell, W. A., Clarke, J. M., and Kochian, L. V. (2005). Zinc effects on cadmium accumulation and partitioning in near-isogenic lines of durum wheat that differ in grain cadmium concentration. New Phytol. 167, 391e401. Hart, J. J., Welch, R. M., Norvell, W. A., and Kochian, L. V. (2002). Transport interactions between cadmium and zinc in roots of bread and durum wheat seedlings. Physiol. Plant. 116, 73e78. Hattori, H., Kuniyasu, H., Chiba, K., and Chino, M. (2006). Effect of chloride application and low soil pH on cadmium uptake from soil by plants. Soil Sci. Plant Nutr. 52, 89e94. Hinesly, T. D., Hansen, L. G., Bray, D. J., and Redborg, K. E. (1985). Transfer of sludgeborne cadmium through plants to chickens. J. Agr. Food Chem. 33, 173e180. Holmgren, G. G. S., Meyer, M. W., Chaney, R. L., and Daniels, R. B. (1993). Cadmium, lead, zinc, copper, and nickel in agricultural soils of the United States of America. J. Environ. Qual. 22, 335e348. Honda, R., Swaddiwudhipong, W., Nishijo, M., Mahasakpan, P., Teeyakasem, W., Ruangyuttikarn, W., Satarug, S., Padungtod, C., and Nakagawa, H. (2010). Cadmium induced renal dysfunction among residents of rice farming area downstream from a zincmineralized belt in Thailand. Toxicol Lett. 198, 26e32. Horiguchi, H., Aoshima, K., Oguma, E., Sasaki, S., Miyamoto, K., Hosoi, Y., Katoh, T., and Kayama, F. (2010). Latest status of cadmium accumulation and its effects on kidneys, bone, and erythropoiesis in inhabitants of the formerly cadmium-polluted Jinzu River Basin in Toyama, Japan, after restoration of rice paddies. Int. Arch. Occup. Environ. Health 83, 953e970. Hornick, S. B., Baker, D. E., and Guss, S. B. (1977). Crop production and animal health problems associated with high soil molybdenum. In W. R. Chappell, and K. K. Peterson (Eds.), Molybdenum in the Environment Vol. 2. (pp. 665e684). New York: Marcel Dekker. Hossain, M., Islam, M. R., Jahiruddin, M., Abedin, A., Islam, S., and Meharg, A. A. (2008). Effects of arsenic-contaminated irrigation water on growth, yield, and nutrient concentration in rice. Commun. Soil Sci. Plant Anal. 39, 302e313. Howe, A., Fung, L. H., Lalor, G., Rattray, R., and Vutchkov, M. (2005). Elemental composition of Jamaican foods 1: A survey of five food crop categories. Environ. Geochem. Health 27, 19e30. Hwang, Y.-H., Bornschein, R. L., Grote, J., Menrath, M., and Roda, S. (1997). Environmental arsenic exposure of children around a former copper smelter site. Environ. Res. 72, 72e81. Ikeda, M., Ezaki, T., Moriguchi, J., Fukui, Y., Okamoto, S., Ukai, H., and Sakurai, H. (2006). No meaningful increase in urinary tubular dysfunction markers in a population with 3 mg cadmium/g creatinine in urine. Biol. Trace Elem. Res. 113, 35e44. Ikeda, M., Ezaki, T., Tsukahara, T., Moriguchi, J., Furuki, K., Fukui, Y., Ukai, H., Okamoto, S., and Sakura, H. (2003). Threshold levels of urinary cadmium in relation to increases in urinary b2-microglobulin among general Japanese populations. Toxicol. Lett. 137, 135e141. Ippolito, J. A., Barbarick, K. A., and Elliott, H. A. (2011). Drinking water treatment residuals: A review of recent uses. J. Environ. Qual. 40, 1e12.

106

Rufus L. Chaney

Isermann, K. (1982). Influence of phosphate fertilization on the cadmium content of the soil determined on the basis of long-term trials in Western Europe (in German). Landwirtschaft. Forsch. Sonderh. 39, 283e301. Jacobs, R. M., Jones, A. O. L., Fry, B. E., Jr., and Fox, M. R. S. (1978). Decreased longterm retention of 115mcadmium in Japanese quail produced by a combined supplement of zinc copper and manganese. J. Nutr. 108, 901e910. Järup, L., and Alfvén, T. (2004). Low level cadmium exposure, renal and bone effects e The OSCAR study. BioMetals 17, 505e509. JECFA ( Joint Expert Committee on Food Aditives). (2010). Summary and Conclusions, 73rd Meeting, Geneva, 8e18 June, 2010. http://www.who.int/foodsafety/publications/chem/ summary73.pdf. Jin, T., Nordberg, G., Ye, T., Bo, M., Wang, H., Zhu, G., Kong, Q., and Bernard, A. (2004). Osteoporosis and renal dysfunction in a general population exposed to cadmium in China. Environ. Res. 96, 353e359. Jones, F. T. (2007). A broad view of arsenic. Poult. Sci. 86, 2e14. Jones, K. C., and Johnston, A. E. (1989). Cadmium in cereal grain and herbage from longterm experimental plots at Rothamsted. Environ. Pollut. 57, 199e216. Jorhem, L., and Slanina, P. (2000). Does organic farming reduce the content of Cd and certain other trace metals in plant foods? A pilot study. J. Sci. Food Agric. 80, 43e48. Jorhem, L., Sundstrom, B., and Engman, J. (2001). Cadmium and other metals in Swedish wheat and rye flours: Longitudinal study, 1983e1997. J AOAC Int. 84, 1984e1992. Juhasz, A. L., Smith, E., Weber, J., Rees, M., Rofe, A., Kuchel, T., Sansom, L., and Naidu, R. (2007). Comparison of in vivo and in vitro methodologies for the assessment of arsenic bioavailability in contaminated soils. Chemosphere 69, 961e966. Jung, J., Isermann, K., and Henjes, G. (1979). The influence of fertilizer phosphates which contain cadmium on the cadmium enrichment of crop soils and useful plants (in German). Landwirtsch. Forsch. 32, 262e274. Keisling, T. C., Lauer, D. A., Walker, M. E., and Henning, R. J. (1977). Visual, tissue, and soil factors associated with Zn toxicity of peanuts. Agron. J. 69, 765e769. Keller, A., and Schulin, R. (2003). Modelling regional-scale mass balances of phosphorus, cadmium and zinc fluxes on arable and dairy farms. Eur. J. Agron. 20, 181e198. Khaokaew, S., Chaney, R. L., Landrot, G., Ginder-Vogel, M., and Sparks, D. L. (2011). Speciation and release kinetics of cadmium in an alkaline paddy soil under various flooding periods and draining conditions. Environ. Sci. Technol. 45, 4249e4255. Khoshgoftarmanesh, A. H., and Chaney, R. L. (2007). Preceding crop affects grain cadmium and zinc of wheat grown in saline soils of central Iran. J. Environ. Qual. 36, 1132e1136. Khoshgoftar, A. H., Shariatmadari, H., Karimian, N., Kalbasi, M., van der Zee, S. E. A. T. M., and Parker, D. R. (2004). Salinity and zinc application effects on phytoavailability of cadmium and zinc. Soil Sci. Soc. Am. J. 68, 1885e1889. Kienholz, E. W., Ward, G. M., Johnson, D. E., Baxter, J., Braude, G., and Stern, G. (1979). Metropolitan Denver sewage sludge fed to feedlot steers. J. Anim. Sci. 48, 735e741. Kim, D.-W., Kim, K.-Y., Choi, B.-S., Youn, P., Ryu, D.-Y., Klaassen, C. E., and Park, C.-D. (2007). Regulation of metal transporters by dietary iron, and the relationship between body iron levels and cadmium uptake. Arch. Toxicol. 81, 327e334. Kim, K.-W., Bang, S., Zhu, Y.-G., Meharg, A. A., and Bhattacharya, P. (2009). Arsenic geochemistry, transport mechanism in the soil-plant system, human and animal health issues. Environ. Int. 35, 453e454.

Food Safety Issues for Mineral and Organic Fertilizers

107

Kirchmann, H., Mattsson, L., and Eriksson, J. (2009). Trace element concentration in wheat grain: Results from the Swedish long-term soil fertility experiments and national monitoring program. Environ. Geochem. Health 31, 561e571. Kjellström, T., Lind, B., Linnman, L., and Elinder, C. G. (1975). Variation of cadmium concentration in Swedish wheat and barley. An indicator of changes in daily intake during the 20th century. Arch. Environ. Health 30, 321e328. Kjellström, T., and Nordberg, G. F. (1978). A kinetic model of cadmium metabolism in the human being. Environ. Res. 16, 248e269. Kobayashi, J. (1978). Pollution by cadmium and the itai-itai disease in Japan. In F. W. Oehme (Ed.), “Toxicity of Heavy Metals in the Environment” (pp. 199e260). New York, NY: Marcel Dekker, Inc. Koeppe, D. E. (1981). Lead: Understanding the minimal toxicity of lead in plants. In N. W. Lepp (Ed.), Effect of Heavy Metal Pollution on Plant Function Vol. 1 (pp. 55e76). London: Applied Science Publ. Kost, D. A., Bigham, J. M., Stehouwer, R. C., Beeghly, J. H., Fowler, R., Traina, S. J., Wolfe, W. E., and Dick, W. A. (2005). Chemical and physical properties of dry flue gas desulfurization products. J. Environ. Qual. 34, 676e686. Kowal, N. E., and Zirkes, M. (1983). Urinary cadmium and b2-microglobulin: Normal values and concentration adjustment. J. Toxicol. Environ. Health 11, 607e624. Kubota, J., and Allaway, W. H. (1972). Geographic distribution of trace element problems. In J. J. Mortvedt, et al. (Eds.), “Micronutrients in Agriculture.” (pp. 525e554). Madison, WI: Soil Science Society of America. Kubota, J., Welch, R. M., and Van Campen, D. (1987). Soil related nutritional problem areas for grazing animals. Adv. Soil Sci. 6, 189e215. Kukier, U., Chaney, R. L., Ryan, J. A., Daniels, W. L., Dowdy, R. H., and Granato, T. C. (2010). Phytoavailability of cadmium in long-term biosolids amended soils. J. Environ. Qual. 39, 519e530. Kuo, S., Huang, B., and Bembenek, R. (2004). The availability to lettuce of zinc and cadmium in a zinc fertilizer. Soil Sci. 169, 363e373. Kuo, S., Huang, B., and Bembenek, R. (2007). Release of cadmium from a triple superphosphate and a phosphate rock in soil. Soil Sci. 172, 257e265. Lalor, G. C. (2008). Review of cadmium transfers from soil to humans and its health effects in the Jamaican environment. Sci. Total Environ. 400, 162e172. Lalor, G. C., Rattray, R., Simpson, R., and Vutchkov, M. (1998). Heavy metals in Jamaica. Part 3: The distribution of cadmium in Jamaican soils. Rev. Int. Contam. Ambien. 14, 7e12. Lalor, G. C., Rattray, R., Williams, N., and Wright, P. (2004). Cadmium levels in kidney and liver of Jamaicans at autopsy. West Ind. Med. J. 53, 76e80. Lamm, S. H., Engel, A., Penn, C. A., Chen, R., and Feinleib, M. (2006). Arsenic cancer risk confounder in southwest Taiwan data set. Environ. Health Perspect. 114, 1077e1082. Lamm, S. H., Engel, A., Kruse, M. B., Feinleib, M., Byrd, D. M., Lai, S., and Wilson, R. (2004). Arsenic in drinking water and bladder cancer mortality in the United States: An analysis based on 133 U.S. counties and 30 years of observation. J. Occup. Environ. Med. 46, 298e306. Langmuir, D. L., Chrostrowski, P., Chaney, R. L., and Vigneault, B. (2005). Issue Paper on Environmental Chemistry of Metals. US-EPA Risk Assessment Forum: Papers Addressing Scientific Issues in the Risk Assessment of Metals. http://cfpub.epa.gov/ncea/cfm/ recordisplay.cfm?deid¼86119. pp. 113. Lasky, T., Sun, W., Kadry, A., and Hoffman, M. K. (2004). Mean total arsenic concentrations in chicken 1989e2000 and estimated exposures for consumers of chicken. Environ. Health Perspect. 112, 18e21.

108

Rufus L. Chaney

Lauria, D. C., Ribeiro, F. C. A., Conti, C. C., and Loureiro, F. A. (2009). Radium and uranium levels in vegetables grown using different farming management systems. J. Environ. Radioact. 100, 176e183. Li, R. L., Stroud, J. R., Ma, J. F., McGrath, S. P., and Zhao, F.-J. (2009). Mitigation of arsenic accumulation in rice with water management and silicon fertilization. Environ. Sci. Technol. 43, 3778e3783. Li, Y.-M., Chaney, R. L., and Schneiter, A. A. (1995). Effect of soil chloride level on cadmium concentration in sunflower kernels. Plant Soil 167, 275e280. Loganathan, P., Hedley, M. J., and Grace, N. D. (2008). Pasture soils contaminated with fertilizer-derived cadmium and fluorine: Livestock effects. Rev. Environ. Contam. Toxicol. 192, 29e66. Loganathan, P., Hedley, M. J., Grace, N. D., Lee, J., Cronin, S. J., Bolan, N. S., and Zanders, M. J. (2003). Fertiliser contaminants in New Zealand grazed pasture with special reference to cadmium and fluorine: A review. Aust. J. Soil Res. 41, 501e532. Loganathan, P., Hedley, M. J., Wallace, G. C., and Roberts, A. H. C. (2001). Fluoride accumulation in pasture forages and soils following long-term applications of phosphorus fertilisers. Environ. Pollut. 115, 275e282. Loganathan, P., Louie, K., Lee, J., Hedley, M. J., Roberts, A. H. C., and Longhurst, R. D. (1999). A model to predict kidney and liver cadmium concentrations in grazing animals. N. Z. J. Agric. Res. 42, 423e432. Lorenz, H., Ocker, H.-D., Brüggemann, J., Weigert, P., and Sonnehorn, M. (1986). Content of cadmium in cereals of the past compared with the present (in German). Eur. Food Res. Technol. 183, 402e405. Lugon-Moulin, N., Ryan, L., Donini, P., and Rossi, L. (2006). Cadmium content of phosphate fertilizers used for tobacco production. Agron. Sustain. Devel. 26, 151e155. Lund, L. J., Betty, E. E., Page, A. L., and Elliott, R. A. (1981). Occurrence of naturally high cadmium levels in soils and its accumulation by vegetation. J. Environ. Qual. 10, 551e556. Lung, S. C., Cheng, H. W., and Fu, C. B. (2008). Potential exposure and risk of fluoride intakes from tea drinks produced in Taiwan. J. Expo. Sci. Environ. Epidemiol. 18, 158e166. Ma, J. F., Yamaji, N., Mitani, N., Xu, X.-Y., Su, Y.-H., McGrath, S. P., and Zhao, F.-J. (2008). Transporters of arsenite in rice and their role in arsenic accumulation in rice grain. Proc. Natl. Acad. Sci. U.S.A. 105, 9931e9935. Malik, M., Chaney, R. L., Brewer, E. P., and Angle, J. S. (2000). Phytoextraction of soil cobalt using hyperaccumulator plants. Int. J. Phytoremed. 2, 319e329. Mason, R. W., Reid, R. N. D., and Brown, P. G. (1989). Superphosphate and its effect on bone fluoride concentrations in sheep. Aust. Vet. J. 66, 120e121. Mays, D. A., and Mortvedt, J. J. (1986). Crop response to soil applications of phosphogypsum. J. Environ. Qual. 15, 78e81. McBride, M. B., and Spiers, G. (2001). Trace element content of selected fertilizers and dairy manures as determined by ICP-MS. Commun. Soil Sci. Plant Anal. 32, 139e156. McDowell, L. R. (2003). “Minerals in Animal and Human Nutrition.” (Second Edition). Amsterdam: Elsevier. McKenna, I. M., and Chaney, R. L. (1995). Characterization of a cadmium-zinc complex in lettuce leaves. Biol. Trace Element Res. 48, 13e29. McKenna, I. M., Chaney, R. L., Tao, S. H., Leach, R. H., Jr., and Williams, F. M. (1992). Interactions of plant zinc and plant species on the bioavailability of plant cadmium to Japanese quail fed lettuce and spinach. Environ. Res. 57, 73e87. McLaughlin, M. J., Bell, M. J., Wright, G. C., and Cozens, G. D. (2000). Uptake and partitioning of cadmium by cultivars of peanut (Arachis hypogaea L.). Plant Soil 222, 51e58.

Food Safety Issues for Mineral and Organic Fertilizers

109

McLaughlin, M. J., Hamon, R. E., McLaren, R. G., Speir, T. W., and Rogers, S. L. (2000). Review: A bioavailability-based rationale for controlling metal and metalloid contamination of agricultural land in Australia and New Zealand. Aust. J. Soil Res. 38, 1037e1086. McLaughlin, M. J., Lambrechts, R. M., Smolders, E., and Smart, M. K. (1998). Effects of sulfate on cadmium uptake by Swiss chard. II. Effects due to sulfate addition to soil. Plant Soil 202, 217e222. McLaughlin, M. J., Maier, N. A., Freeman, K., Tiller, K. G., Williams, C. M. J., and Smart, M. K. (1995). Effect of potassic and phosphatic fertilizer type, fertilizer Cd concentration and zinc rate on cadmium uptake by potatoes. Fert. Res. 40, 63e70. McLaughlin, M. J., Maier, N. A., Smart, M. K., Sparrow, L. A., and McKay, A. (1999). Prediction of cadmium concentrations in potato tubers (Solanum tuberosum L.) by pre-plant soil and irrigation water analyses. Aust. J. Soil Res. 37, 191e207. McLaughlin, M. J., Palmer, L. T., Tiller, K. G., Beech, T. W., and Smart, M. K. (1994). Increasing soil salinity causes elevated cadmium concentrations in field-grown potato tubers. J. Environ. Qual. 23, 1013e1018. McLaughlin, M. J., Parker, D. R., and Clarke, J. M. (1999). Metals and micronutrients e Food safety issues. Field Crops Res. 60, 143e163. McLaughlin, M. J., Simpson, P. G., Fleming, N., Stevens, D. P., Cozens, G., and Smart, M. K. (1997). Effect of fertilizer type on cadmium and fluorine concentrations in clover herbage. Aust. J. Exp. Agric. 37, 1019e1026. McLaughlin, M. J., Stevens, D. P., Keerthisinghe, D. G., Cayley, J. W. D., and Ridley, A. M. (2001). Contamination of soil with fluoride by long-term application of superphosphates to pastures and risk to grazing animals. Aust. J. Soil Res. 39, 627e640. McLaughlin, M. J., Tiller, K. G., Naidu, R., and Stevens, D. P. (1996). Review: The behaviour and environmental impact of contaminants in fertilizers. Aust. J. Soil Res. 34, 1e54. Meharg, A. A., and Raab, A. (2010). Getting to the bottom of arsenic standards and guidelines. Environ. Sci. Technol. 44, 4395e4399. Meliker, J. R., Slotnick, M. J., AvRuskin, G. A., Schottenfeld, D., Jacquez, G. M., Wilson, M. L., Goovaerts, P., Franzblau, A., and Nriagu, J. O. (2010). Lifetime exposure to arsenic in drinking water and bladder cancer: A population-based case-control study in Michigan, USA. Cancer Causes Contr. 21, 745e757. Mellum, H. K., Arnesen, A. K. M., and Singh, B. R. (1998). Extractability and plant uptake of heavy metals in alum shale soils. Commun. Soil Sci. Plant Anal. 29, 1183e1198. Mench, M. J. (1998). Cadmium availability to plants in relation to major long-term changes in agronomy systems. Agric. Ecosyst. Environ. 67, 175e187. Menzel, R. G. (1968). Uranium, radium, and thorium content in phosphate rocks and their possible radiation hazard. J. Agr. Food Chem. 16, 231e234. Merwin, I., Pruyne, P. T., Ebel, G. J., Jr., Manzell, K. L., and Lisk, D. J. (1994). Persistence, phytotoxicity, and management of arsenic, lead and mercury residues in old orchard soils of New York State. Chemosphere 29, 1361e1367. Meunier, L., Walker, S. R., Wragg, J., Parsons, M. B., Koch, I., Jamieson, H. E., and Reimer, K. J. (2010). Effects of soil composition and mineralogy on the bioaccessibility of arsenic from tailings and soil in gold mine districts of Nova Scotia. Environ. Sci. Technol. 44, 2667e2674. Michaud, A. M., Bravin, M. N., Galleguillos, M., and Hinsinger, P. (2007). Copper uptake and phytotoxicity as assessed in situ for durum wheat (Triticum turgidum durum L.) cultivated in Cu-contaminated, former vineyard soils. Plant Soil 298, 99e111. Mielke, H. W., Anderson, J. C., Berry, K. W., Mielke, P. W., Chaney, R. L., and Leech, R. L. (1983). Lead concentrations in inner city soils as a factor in the child lead problem. Am. J. Public Health 73, 1366e1369.

110

Rufus L. Chaney

Milhaud, G., Clauw, M., and Joseph-Enriquez, B. (1989). Bioavailability in soil fluoride in sheep. Fluoride 22, 188e194. Million, J. B., Sartain, J. B., Gonzalez, R. X., and Carrier, W. D., III (1994). Radium-226 and calcium uptake by crops grown in mixtures of sand and clay tailings from phosphate mining. J. Environ. Qual. 23, 671e676. Milvy, P., and Cothern, C. R. (1989). Naturally occurring radionuclides in drinking water an exercise in risk benefit analysis. Environ. Geochem. Health 11, 63e72. Mishima, S.-I., Kimura, R., and Inoue, T. (2004). Estimation of cadmium load on Japanese farmland associated with the application of chemical fertilizers and livestock excreta. Soil Sci. Plant Nutr. 50, 263e267. Moberg, A., Hallmans, G., Sjöström, R., and Wing, K. R. (1987). The effect of wheat bran on the absorption and accumulation of cadmium in rats. Br. J. Nutr. 58, 383e391. Mortvedt, J. J. (1984). Cadmium and zinc uptake by vegetable tissues after nine annual applications of phosphate fertilizer to soils. Trace Subst. Environ. Health 18, 283e291. Mortvedt, J. J. (1985). Plant uptake of heavy metals in zinc fertilizers made from industrial by-products. J. Environ. Qual. 14, 424e427. Mortvedt, J. J. (1987). Cadmium levels in soils and plants from some long-term soil fertility experiments in the United States of America. J. Environ. Qual. 16, 137e142. Mortvedt, J. J. (1994). Plant and soil relationships of uranium and thorium decay series radionuclides e A review. J. Environ. Qual. 23, 643e650. Mortvedt, J. J. (1996). Heavy metal contaminants in inorganic and organic fertilizers. Fert. Res. 43, 55e61. Mortvedt, J. J., Mays, D. A., and Osborn, G. (1981). Uptake by wheat of cadmium and other heavy metal contaminants in phosphate fertilizers. J. Environ. Qual. 10, 193e197. Mulla, D. J., Page, A. L., and Gange, T. J. (1980). Cadmium accumulations and bioavailability in soils from long-term phosphorus fertilization. J. Environ. Qual. 9, 408e412. Murakami, M., Nakagawa, F., Ae, N., Ito, M., and Arao, T. (2009). Phytoextraction by rice capable of accumulating Cd at high levels: Reduction of cd content of rice grain. Environ. Sci. Technol. 43, 5878e5883. Nakamura, Y., Narukawa, T., and Yoshinaga, J. (2008). Cancer risk to Japanese population from the consumption of inorganic arsenic in cooked hijiki. J. Agr. Food Chem. 56, 2536e2540. Nicholson, F. A., Chambers, B. J., Williams, J. R., and Unwin, R. J. (1999). Heavy metal contents of livestock feeds and animal manures in England and Wales. Bioresour. Technol. 70, 23e31. Nicholson, F. A., Smith, S. R., Alloway, B. J., Carlton-Smith, C., and Chambers, B. J. (2003). An inventory of heavy metals inputs to agricultural soils in England and Wales. Sci. Total Environ. 311, 205e219. Nogawa, K., Kobayashi, E., Okubo, Y., and Suwazono, Y. (2004). Environmental cadmium exposure, adverse effects and preventive measures in Japan. BioMetals 17, 581e587. Norvell, W. A., Wu, J., Hopkins, D. J., and Welch, R. M. (2000). Association of cadmium in durum wheat grain with soil chloride and chelate-extractable soil cadmium. Soil Sci. Soc. Am. J. 64, 2162e2168. NRC (National Research Council). (1980). Mineral Tolerance of Domestic Animals. Washington, DC: National Academy of Sciences. Nziguheba, G., and Smolders, E. (2008). Inputs of trace elements in agricultural soils via phosphate fertilizers in European countries. Sci Total Environ. 390, 53e57. O'Connor, G. A., Brobst, R. B., Chaney, R. L., Kincaid, R.l., McDowell, L. R., Pierzynski, G. M., Rubin, A., and Van Riper, G. G. (2001). Molybdenum standards for the land application of biosolids. J. Environ. Qual. 30, 1490e1507.

Food Safety Issues for Mineral and Organic Fertilizers

111

Ohlendorf, H. M., Oldfield, J. E., Sarka, M. K., and Aldrich, T. W. (1986). Embryonic mortality and abnormalities of aquatic birds: Apparent impacts by selenium from irrigation drain water. Sci. Total Environ. 52, 49e63. Oliver, D. P., Hannam, R., Tiller, K. G., Wilhelm, N. S., and Merry, R. H. (1994). The effects of zinc fertilization on cadmium concentration in wheat grain. J. Environ. Qual. 23, 705e711. Oliver, D. P., Schultz, J. E., Tiller, K. G., and Merry, R. H. (1993). The effect of crop rotations and tillage practices on cadmium concentrations in wheat grain. Aust. J. Agric. Res. 44, 1221e1234. Olsson, I.-M., Bensryd, I., Lundh, T., Ottosson, H., Skerfving, S., and Oskarsson, A. (2002). Cadmium in blood and urineeImpact of sex, age, dietary intake, iron status, and former smokingeAssociation of renal effects. Environ. Health Perspect. 110, 1185e1190. Park, J. D., Cherrington, N. J., and Klaassen, C. D. (2002). Intestinal absorption of cadmium is associated with divalent metal transporter 1 in rats. Toxicol. Sci. 68, 288e294. Peryea, F. J., and Creger, T. L. (1994). Vertical distribution of lead and arsenic in soils contaminated with lead arsenate pesticide residues. Water Air Soil Pollut. 78, 297e306. Phillips, R. L., and Meyer, R. D. (1993). Molybdenum concentration of alfalfa in Kern County, California: 1950 versus 1985. Commun. Soil Sci. Plant Anal. 24, 2725e1731. Pierzynski, G. M., and Jacobs, L. W. (1986). Molybdenum accumulation by corn and soybeans from a molybdenum-rich sewage sludge. J. Environ. Qual. 15, 394e398. Rahman, F. A., Allan, D. L., Rosen, C. J., and Sadowsky, M. J. (2004). Arsenic availability from chromated copper arsenate (CCA)-treated wood. J. Environ. Qual. 33, 173e180. Rahman, M. A., Hasegawa, H., Rahman, M. M., Miah, M. A. M., and Tasmin, A. (2008). Straighthead disease of rice (Oryza sativa L.) induced by arsenic toxicity. Environ. Exp. Bot. 62, 54e59. Raisbeck, M. F. (1982). Is polioencephalomalacia associated with high-sulfate diets? J. Am. Vet. Med. Assoc. 180, 1303e1305. Ransome, L. S., and Dowdy, R. H. (1987). Soybean growth and boron distribution in a sandy soil amended with scrubber sludge. J. Environ. Qual. 16, 171e175. Rea, R. E. (1979). A rapid method for determination of fluoride in sewage sludges. Wat. Pollut. Contr. 78, 139e142. Reeves, P. G., and Chaney, R. L. (2001). Mineral nutrient status of female rats affects the absorption and organ distribution of cadmium from sunflower kernels (Helianthus annuus L.). Environ. Res. 85, 215e225. Reeves, P. G., and Chaney, R. L. (2002). Nutritional status affects the absorption and whole-body and organ retention of cadmium in rats fed rice-based diets. Environ. Sci. Technol. 36, 2684e2692. Reeves, P. G., and Chaney, R. L. (2004). Marginal nutritional status of zinc, iron, and calcium increases cadmium retention in the duodenum and other organs of rats fed a rice-based diet. Environ. Res. 96, 311e322. Reeves, P. G., Chaney, R. L., Simmons, R. W., and Cherian, M. G. (2005). Metallothionein induction is not involved in cadmium accumulation in the duodenum of mice and rats fed diets containing high-cadmium-rice or sunflower kernels and a marginal supply of zinc, iron, and calcium. J. Nutr. 135, 99e108. Reeves, P. G., and Chaney, R. L. (2008). Bioavailability as an issue in risk assessment and management of food cadmium: A review. Sci. Total Environ. 398, 13e19. Reeves, P. G., Nielsen, E. J., O'Brien-Nimens, C., and Vanderpool, R. A. (2001). Cadmium bioavailability from edible sunflower kernels: A long-term study with men and women volunteers. Environ. Res. A87, 81e91. Reuss, J. O., Dooley, H. L., and Griffis, W. (1978). Uptake of cadmium from phosphate fertilizers by peas, radishes, and lettuce. J. Environ. Qual. 7, 128e133.

112

Rufus L. Chaney

Reuther, W., and Smith, P. F. (1954). Toxic effects of accumulated copper in Florida soils. Soil Sci. Soc. Fla. Proc. 14, 17e23. Richards, I. R., Clayton, C. J., and Reeve, A. J. K. (1988). Effects of long-term fertilizer phosphorus application on soil and crop phosphorus and cadmium contents. J. Agric. Sci. 131, 187e195. Roberts, A. H. C., Longhurst, R. D., and Brown, M. W. (1994). Cadmium status of soils, plants, and grazing animals in New Zealand. N. Z. J. Agric. Res. 37, 119e129. Roberts, S. M., Munson, J. W., Lowney, Y. W., and Ruby, M. V. (2007). Relative oral bioavailability of arsenic from contaminated soils measured in the cynomolgus monkey. Toxicol. Sci. 95, 281e288. Roberts, S. M., Weimar, W. R., Vinson, J. R. T., Munson, J. W., and Bergeron, R. J. (2002). Measurement of arsenic bioavailability in soil using a primate model. Toxicol. Sci. 67, 303e310. Rodriguez, R. R., Basta, N. T., Casteel, S. W., Armstrong, F. P., and Ward, D. C. (2003). Chemical extraction methods to assess bioavailable arsenic in soil and solid media. J. Environ. Qual. 32, 876e884. Rothbaum, H. P., McGaveston, D. A., Wall, T., Johnston, A. E., and Mattingly, G. E. G. (1979). Uranium accumulation in soils from long continued applications of super phosphate. J. Soil Sci. 30, 147e154. Rutherford, P. M., Dudas, M. J., and Samek, R. A. (1994). Environmental impacts of phosphogypsum. Sci. Total Environ. 149, 1e38. Ryan, J. A., Berti, W. R., Brown, S. L., Casteel, S. W., Chaney, R. L., Doolan, M., Grevatt, P., Hallfrisch, J. G., Maddaloni, M., and Mosby, D. (2004). Reducing children's risk from soil lead: Summary of a field experiment. Environ. Sci. Technol. 38, 18Ae24A. Sarasua, S. M., McGeehin, M. A., Stallings, F. L., Terracciano, G. J., Amler, R. W., Logue, J. N., and Fox, J. M. (1995). Final Report. Technical Assistance to the Pennsylvania Department of Health. Biologic Indicators of Exposure to Cadmium and Lead Palmerton, PA. Part II. (May 1995). pp. 57. Agency for Toxic Substances and Disease Registry. Atlanta, GA: US-DHHS. Sauerbeck, D. (1982). Environmental significance of the cadmium content in phosphorus fertilizers (in German). Landbauforsch. Volk 32, 192e197. Scheckel, K. G., Chaney, R. L., Basta, N. T., and Ryan, J. A. (2009). Advances in assessing bioavailability of metal(loid)s in contaminated soils. Adv. Agron. 104, 1e52. Scheckel, K. G., and Ryan, J. A. (2004). Spectroscopic speciation and quantification of lead in phosphate-amended soils. J. Environ. Qual. 33, 1288e1295. Scheidegger, A. M., Lamble, G. M., and Sparks, D. L. (1997). Spectroscopic evidence for the formation of mixed-cation hydroxide phases upon metal sorption on clays and aluminum oxides. J. Colloid Interfac. Sci. 186, 118e128. Schroder, J. L., Basta, N. T., Payton, M., Wilson, J. A., Carlson, R. I., Janz, D. M., and Lochmiller, R. L. (2003). Ecotoxicological risks associated with land treatment of petrochemical wastes. I. Residual soil contamination and bioaccumulation by cotton rats (Sigmodon hispidus). J. Toxicol. Environ. Health A66, 305e325. Schroder, J. L., Basta, N. T., Rafferty, D. P., Lochmiller, R. L., Kim, S., Qualls, C. W., Jr., and Mcbee, K. (1999). Soil and vegetation fluoride exposure pathways to cotton rats on a petrochemical-contaminated landfarm. Environ. Toxicol. Chem. 18, 2023e2033. Sharma, R. P., Kjellström, T., and McKenzie, J. M. (1983). Cadmium in blood and urine among smokers and non-smokers with high cadmium intake via food. Toxicol. 29, 163e171. Sheppard, S. C., Evenden, W. G., and Pollock, R. J. (1989). Uptake of natural radionuclides by field and garden crops. Can. J. Soil Sci. 69, 751e767.

Food Safety Issues for Mineral and Organic Fertilizers

113

Sheppard, S. C., Grant, C. A., Sheppard, M. I., de Jong, R., and Long, J. (2009). Risk indicator for agricultural inputs of trace elements to Canadian soils. J. Environ. Qual. 38, 919e932. Shupe, J. L., Miner, M. L., Harris, L. F., and Greenwood, D. A. (1962). Relative effects of feeding hay atmospherically contaminated by fluoride residue, normal hay plus calcium fluoride, and normal hay plus sodium fluoride to dairy heifers. Am. J. Vet. Res. 23, 777e787. Simmons, R. W., Noble, A. D., Pongsakul, P., Sukreeyapongse, O., and Chinabut, N. (2008). Analysis of field-moist Cd contaminated paddy soils during rice grain fill allows reliable prediction of grain Cd levels. Plant Soil 302, 125e137. Simmons, R. W., Noble, A. D., Pongsakul, P., Sukreeyapongse, O., and Chinabut, N. (2009). Cadmium-hazard mapping using a general linear regression model (Irr-Cad) for rapid risk assessment. Environ. Geochem. Health 31, 71e79. Simmons, R. W., Pongsakul, P., Chaney, R. L., Saiyasitpanich, D., Klinphoklap, S., and Nobuntou, W. (2003). The relative exclusion of zinc and iron from rice grain in relation to rice grain cadmium as compared to soybean: Implications for human health. Plant. Soil 257, 163e170. Simmons, R. W., Pongsakul, P., Saiyasitpanich, D., and Klinphoklap, S. (2005). Elevated levels of cadmium and zinc in paddy soils and elevated levels of cadmium in rice grain downstream of a zinc mineralized area in Thailand: Implications for public health. Environ. Geochem. Health 27, 501e511. Singh, B. R. (1990). Cadmium and fluoride uptake by oats and rape from phosphate fertilizers in two different soils. Norw. J. Agric. Sci. 4, 239e249. Singh, B. R., and Myhr, K. (1998). Cadmium uptake by barley as affected by Cd sources and pH levels. Geoderma 84, 185e194. Sirot, V., Samieri, C., Volatier, J.-L., and LeBlanc, J.-C. (2008). Cadmium dietary intake and biomarker data in French high seafood consumers. J. Expos. Sci. Environ. Epidemiol. 28, 400e409. Smilde, K. W., and van Luit, B. (1983). The Effect of Phosphate Fertilizer Cadmium on Cadmium in Soils and Crops. Rapport Institut Bodemvruchtbaarheid No. 6-83:1e17. Netherlands: Haren. Smith, D. B., Cannon, W. F., Woodruff, L. G., Garrett, R. G., Klassen, R. K., Kilburn, J. E., Horton, J. D., King, H. D., Goldhaber, M. B., and Morrison, J. M. (2005). “Major- and Trace-element Concentrations in Soils from Two Continental-scale Transects of the United States and Canada.” USGS Open-File Report 2005e1253. Smith, R. M., Leach, R. M., Muller, L. D., Griel, L. C., Jr., and Baker, D. E. (1991). Effects of long-term dietary cadmium chloride on tissue, milk, and urine mineral concentrations of lactating dairy cows. J. Anim. Sci. 69, 4088e4096. Smolders, E., Buekers, J., Oliver, I., and Mclaughlin, M. J. (2004). Soil properties affecting toxicity of zinc to soil microbial properties in laboratory-spiked and field-contaminated soils. Environ. Toxicol. Chem. 23, 2633e2640. Smolders, E., and McLaughlin, J. M. (1996a). Effect of Cl on Cd uptake by Swiss chard in unbuffered and chelate-buffered nutrient solutions. Plant Soil 179, 57e64. Smolders, E., and McLaughlin, J. M. (1996b). Chloride increases Cd uptake in Swiss chard in a resin-buffered nutrient solution. Soil Sci. Soc. Am. J. 60, 1443e1447. Smolders, E., Oorts, K., van Sprang, P., Schoeters, I., Janssen, C. R., McGrath, S. P., and McLaughlin, M. J. (2009). Toxicity of trace metals in soils as affected by soil type and aging after contamination: Using calibrated bioavailability models to set ecological soil standards. Environ. Toxicol. Chem. 28, 1633e1642. Sommers, L. E. (1977). Chemical composition of sewage sludges and analysis of their potential use as fertilizers. J. Environ. Qual. 6, 225e232.

114

Rufus L. Chaney

Sparrow, L. A., Salardini, A. A., and Bishop, A. C. (1993). Field studies on cadmium in potatoes (Solanum tuberosum L.). II. Response of cv. Russet Burbank and Kennebec to two double superphosphates of different cadmium concentrations. Aust. J. Agric. Res. 44, 855e861. Sparrow, L. A., Salardini, A. A., and Johnsone, J. (1994). Field studies on cadmium in potatoes (Solanum tuberosum L.). III. Response of cv. Russet Burbank to sources of banded potassium. Aust. J. Agric. Res. 45, 243e249. Speir, T. W., Kettles, H. A., Percival, H. J., and Parshotam, A. (1999). Is soil acidification the cause of biochemical responses when soils are amended with heavy metal salts? Soil Biol. Biochem. 31, 1953e1961. Staed, J. B., Miller, D. M., Brye, K. R., Daniel, T. C., Rom, C., and Gbur, E. E. (2009). Land use effects on near-surface soil arsenic in the Ozark highlands. Soil Sci. 174, 121e129. Stanek, E. J., III, Calabrese, E. J., and Zorn, M. (2001). Soil ingestion distributions for Monte Carlo risk assessment in children. Human Ecol. Risk Assess. 7, 357e368. Stehouwer, R. C., Wolf, A. M., and Doty, W. T. (2000). Chemical monitoring of sewage sludge in Pennsylvania: Variability and application uncertainty. J. Environ. Qual. 29, 1686e1695. Sterrett, S. B., Chaney, R. L., Hirsch, C. E., and Mielke, H. W. (1996). Influence of amendments on yield and heavy metal accumulation of lettuce grown in urban garden soils. Environ. Geochem. Health 18, 135e142. Stevens, D. P., McLaughlin, M. J., Randall, P. J., and Keerthisinghe, G. (2000). Effect of fluoride supply on fluoride concentrations in five pasture species: Levels required to reach phytotoxic or potentially zootoxic concentrations in plant tissue. Plant Soil 227, 223e233. Stewart, D. J., Manley, T. R., White, D. A., and Harrison, D. L. (1974). Fluorine residues on pasture, in soil, and in sheep urine, resulting from topdressing with superphosphate. N. Z. J. Exp. Agric. 2, 129e133. Strehlow, C. D., and Barltrop, D. (1988). The Shipham Report e An investigation into cadmium concentrations and its implications for human health: 6. Health studies. Sci. Total Environ. 75, 101e133. Stroud, J. L., Li, H. F., Lopez-Bellido, F. J., Broadley, M. R., Foot, I., FairweatherTait, S. J., Hart, D. J., Hurst, R., Knott, P., Mowat, H., Norman, K., Scott, P., Tucker, M., White, P. J., McGrath, S. P., and Zhao, F.-J. (2010). Impact of sulphur fertilisation on crop response to selenium fertilisation. Plant Soil 332, 31e40. Stuczynski, T., Siebielec, G., Daniels, W. L., McCarty, G., and Chaney, R. L. (2007). Biological aspects of metal waste reclamation with biosolids. J. Environ. Qual. 36, 1154e1162. Suttie, J. W. (1980). Nutritional aspects of fluoride toxicosis. J. Anim. Sci. 51, 759e766. Suttle, N. F., Abrahams, P., and Thornton, I. (1984). The role of a soil  dietary sulphur interaction in the impairment of copper absorption by ingested soil in sheep. J. Agr. Sci. 103, 81e86. Suwazono, Y., Sand, S., Vahter, M., Filipsson, A. F., Skerfving, S., Lidfeldt, J., and Åkesson, A. (2006). Benchmark dose for cadmium-induced renal effects in humans. Environ. Health Perspect. 114, 1072e1076. Taheri, S., Khoshgoftarmanesh, A. H., Shariatmadari, H., and Chaney, R. L. (2011). Kinetics of zinc release from ground tire rubber and rubber ash in a calcareous soil as alternatives to Zn fertilizers. Plant Soil 341, 89e97. Takijima, Y., and Katsumi, F. (1973). Cadmium contamination of soils and rice plants caused by zinc mining. l. Production of high-cadmium rice on the paddy fields in lower reaches of the mine station. Soil Sci. Plant Nutr. 19, 29e38.

Food Safety Issues for Mineral and Organic Fertilizers

115

Tohyama, C., Shaikh, Z. A., Nogawa, K., Kobayashi, E., and Honda, R. (1981). Elevated urinary excretion of metallothionein due to environmental cadmium exposure. Toxicology 20, 289e297. Tracy, B. L., Prantl, F. A., and Quinn, J. M. (1983). Transfer of 226Ra, 210Pb and uranium from soil to garden produce. Assessment of risk. Health Phys. 44, 469e477. Tsuji, J. S., Van Kerkhove, M. D., Kaetzel, R. S., Scrafford, C. G., Mink, P. J., Barraj, L. M., Crecelius, E. A., and Goodman, M. (2005). Evaluation of exposure to arsenic in residential soil. Environ. Health Perspect. 113, 1735e1740. Tsuji, J. S., Yost, L. J., Barraj, L. M., Scrafford, C. G., and Mink, P. J. (2007). Use of background inorganic arsenic exposures to provide perspective on risk assessment results. Reg. Toxicol. Pharmacol. 48, 59e68. Underwood, E. J. (1977). “Trace Elements in Human and Animal Nutrition.” (4th ed.) New York: Academic Press. pp. 545. US-EPA (Environmental Protection Agency). (1995). A Guide to the Biosolids Risk Assessments for the EPA Part 503 Rule. Washington, DC: US-EPA, Office of Wastewater Management. http://water.epa.gov/scitech/wastetech/biosolids/503rule_index.cfm EPA/832-B93e005. US-EPA (Environmental Protection Agency). (2002). Zinc Fertilizers made from Recycled Hazardous Secondary Materials. Fed. Reg. 67, 48393e48415. http://www.epa.gov/ osw/hazard/recycling/fertiliz/fert-fr.pdf. US-EPA (Environmental Protection Agency). (2009). Targeted National Sewage Sludge Survey: Overview. Report. No. EPA 822-R-08e014. Washington, DC: US Environmental Protection Agency, Office of Water. http://water.epa.gov/scitech/wastetech/ biosolids/tnsss-overview.cfm. US-EPA (Environmental Protection Agency). (2010). Toxicological Review of Inorganic Arsenic. EPA/635/R-10/001. Washington, DC: US Environmental Protection Agency. Vahter, M., Berglund, M., Nermell, B., and Åkesson, A. (1996). Bioavailability of cadmium from shellfish and mixed diet in women. Toxicol. Appl. Pharmacol. 136, 332e341. Vahter, M., Marafante, E., and Dencker, L. (1983). Metabolism of arsenobetaine in mice, rats and rabbits. Sci. Total Environ. 30, 197e211. Voegelin, A., Pfister, S., Scheinost, A. C., Marcus, M. A., and Kretzschmar, R. (2005). Changes in zinc speciation in field soil after contamination with zinc oxide. Environ. Sci. Technol. 39, 6616e6623. Voegelin, A., Scheinost, A. C., Bühlmann, K., Barmettler, K., and Kretzschmar, R. (2002). Slow formation and dissolution of Zn precipitates in soil: A combined column-transport and XAFS study. Environ. Sci. Technol. 36, 3749e3754. Wakefield, Z. T. (1980). Distribution of Cadmium and Selected Heavy Metals in Phosphate Fertilizer Processing Bull. Y-159. Muscle Shoals, AL: National Fertilizer Development Center Tennessee Valley Authority. Wångstrand, H., Eriksson, J., and Öborn, I. (2007). Cadmium concentration in winter wheat as affected by nitrogen fertilization. Eur. J. Agron. 26, 209e214. Weggler-Beaton, K., McLaughlin, M. J., and Graham, R. D. (2000). Salinity increases cadmium uptake by wheat and Swiss chard from soil amended with biosolids. Aust. J. Soil Res. 38, 37e46. Weinstein, L. H. (1977). Fluoride and plant life. J. Occup. Med. 19, 49e78. Weinstein, L. H., and Davison, A. D. (2003). Native plant species suitable as bioindicators and biomonitors for airborne fluoride. Environ. Pollut. 125, 3e11. Westfall, D. G., Mortvedt, J. J., Peterson, G. A., and Gangloff, W. J. (2005). Efficient and environmentally safe use of micronutrients in agriculture. Commun. Soil Sci. Plant Anal. 36, 169e182.

116

Rufus L. Chaney

White, M. C., Decker, A. M., and Chaney, R. L. (1979). Differential cultivar tolerance of soybean to phytotoxic levels of soil Zn. I. Range of cultivar response. Agron. J. 71, 121e126. Williams, A. G., Scheckel, K. G., Tolaymat, T., and Impellitteri, C. A. (2006). Mineralogy and characterization of arsenic, iron, and lead in a mine waste-derived fertilizer. Environ. Sci. Technol. 40, 4874e4879. Williams, C. H., and David, D. J. (1973). The effect of superphosphate on the cadmium content of soils and plants. Aust. J. Soil Res. 11, 43e56. Wilson, D. (2001). Fateful Harvest: The True Story of a Small Town, a Global Industry and a Toxic Secret. New York: Harper Collins. Wing, A. M. (1993). The effects of whole wheat, wheat bran and zinc in the diet on the absorption and accumulation of cadmium in rats. Brit. J. Nutr. 69, 199e209. Wing, A. M., Wing, K., Tidehag, P., Hallmans, G., and Sjöström, R. (1992). Cadmium accumulation from diets with and without wheat bran in rats with different iron status. Nutr. Res. 12, 1205e1215. Wixson, B. G., Gale, N. L., and Davies, B. E. (1984). Further characterization and use of tailings and chat from Missouri's old lead belt as agricultural lime. Trace Subst. Environ. Health 18, 251e261. Wolnik, K. A., Fricke, F. L., Capar, S. G., Braude, G. L., Meyer, M. W., Satzger, R. D., and Bonnin, E. (1983). Elements in major raw agricultural crops in the United States. l. Cadmium and lead in lettuce, peanuts, potatoes, soybeans, sweet corn, and wheat. J. Agr. Food Chem. 31, 1240e1244. Wright, P. R. D., Rattray, R., Lalor, G., and Hanson, R. (2010). Minimal health impact from exposure to diet-sourced cadmium on a population in central Jamaica. Environ. Geochem. Health 32, 567e581. Wu, J., Norvell, W. A., Hopkins, D. G., and Welch, R. M. (2002). Spatial variability of grain cadmium and soil characteristics in a durum wheat field. Soil Sci. Soc. Am. J. 66, 268e275. Yang, G., Wang, S., Zhou, R., and Sun, S. (1983). Endemic selenium intoxication of humans in China. Am. J. Clin. Nutr. 37, 872e881. Zhao, F.-J., Ma, J. F., Meharg, A. A., and McGrath, S. P. (2009). Arsenic uptake and metabolism in plants. New Phytol. 181, 777e794. Zhao, F.-J., McGrath, S. P., and Meharg, A. A. (2010). Arsenic as a food chain contaminant: Mechanisms of plant uptake and metabolism and mitigation strategies. Annu. Rev. Plant Biol. 61, 535e559. Zia, M. H., Codling, E. E., Scheckel, K. G., and Chaney, R. L. (2011). In-vitro and In-vivo approaches for the measurement of oral bioavailability of lead (Pb) in soil: A critical review. Environ. Pollut. 159, 2320e2327.